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IC 9183 



Bureau of Mines Information Circular/1988 



AP R 






Mine Drainage and 
Surface Mine Reclamation 

Proceedings of a conference sponsored by The American 
Society for Surface Mining and Reclamation, The Bureau 
of Mines, and The Office of Surface Mining Reclamation 
and Enforcement. The conference was held in 
Pittsburgh, Pennsylvania on April 19-21, 1988. 

Volume I: Mine Water and Mine Waste. 




UNITED STATES DEPARTMENT OF THE INTERIOR 




Information Circular 9183 



Mine Drainage and 
Surface Mine Reclamation 

Proceedings of a conference sponsored by The American 
Society for Surface Mining and Reclamation, The Bureau 
of Mines, and The Office of Surface Mining Reclamation 
and Enforcement. The conference was held in 
Pittsburgh, Pennsylvania on April 19-21, 1988. 

Volume I: Mine Water and Mine Waste. 




*'*c'TT* 



Volume II: Mine Reclamation, Abandoned Mine 
Lands and Policy Issues, is also available 
from The Bureau of Mines as Information 
Circular 9184. 



UNITED STATES DEPARTMENT OF THE INTERIOR 
Donald Paul Hodel, Secretary 

BUREAU OF MINES 

David S. Brown, Acting Director 



^ 






QS-£ol7i3 



UNIT OF MEASURE ABBREVIATIONS USED IN THIS REPORT 

With Factors for Conversion to U.S. Customary Units 
and the International System of Units (SI)'' 



Abbreviation 


Unit of measure 


To convert to — 


Multiply by — 


or unit 








acre 


acre 


hectares 


0.405 


bar 


bar 


NA 




bu 


bushel 


NA 




C 


coulomb 


NA 




cm 


centimeter 


inches 


0.3937 


eV 


electron volt 


NA 




ft 


foot 


meters 


0.3048 


ft2 


square foot 


square centimeter 


929.0 


ft3 


cubic foot 


cubic meters 


0.028 


ft/s 


foot per second 


centimeters per second 


30.48 


g 


gram 


ounces 


0.0353 


gal 


gallon 


liters 


3.785 


gal/h 


gallon per hour 


liters per hour 


3.785 


gal/min 


gallon per minute 


liters per minute 


3.785 


h 


hour 


NA 




ha 


hectare 


acres 


2.471 


in 


inch 


centimeters 


2.54 


K 


kelvin 


NA 




kg 


kilogram 


pounds 


2.205 


L 


liter 


cubic inches 


61 .025 


lb 


pound 


kilograms 


0.4536 


lb/min 


pounds per minute 


kilograms per minute 


0.4536 


L/min 


liter per minute 


gallons per minute 


0.2642 


m 


meter 


feet 


3.28 


M 


mole per liter 






M 


mega 


NA 




m3 


cubic meter 


cubic yards 


1.308 


m3/s 


cubic meter per second 


gallons per second 


264.2 


mil 


mil 


inch 


0.001 inch 


mile 


mile 


kilometers 


1.609 


mg 


milligram 


grains 


0.0154 


mg/L 


milligram per liter 


NA 




mg/(L-h) 


milligram per liter per hour 


NA 




min 


minute 


NA 




mL 


milliliter 


cubic inches 


0.061 


mm 


millimeter 


inches 


0.0394 


mmho/m 


millimho per meter 


NA 




mol .. 

mol/L 

mv 

ym 


iifi?sv t liter 

micrometer 


inches 


3.94 x 10"5 


ohm 


ohm 


NA 




pet 


percent 


NA 




psi 


pound per square inch 


grams per square 
centimeter 


70.307 


std ft3/min 


standard cubic foot per 


NA 






minute 


NA 




ton 


ton 


metric tons 


0.907 


yr 


year 


NA 





NA Not applicable. 

1 0wing to the preference of individual authors, U.S. customary and SI units have both 
been used in this report. Conversion factors are provided for the assistance of the 
reader. 



TABLE OF CONTENTS 



Acid Mine Drainage: Prediction 

Page 

Use of Acid-Base Accounts in Premining Prediction of Acid Drainage Potential: A New Approach 

from Northern West Virginia 2 

R. S. diPretoro and H. W. Rauch 

Evaluation of Overburden Analytical Methods as Means to Predict Post-Mining Coal Mine Drainage Quality.. 11 

P. M. Erickson and R. S. Hedin 

Relationships Between the Initial Geochemistry and Leachate Chemistry of Weathering Overburden Samples.. 21 

R. S. Hedin and P. M. Erickson 

Application of Acid-Base Analysis to Wastes From Base Metal and Precious Metal Mines 29 

S. D. Miller and G. S. Murray 

A Study of Mine Drainage Quality and Prediction Using Overburden Analysis and Paleoenvironmental 
Reconstructions, Fayette County, Pennsylvania 33 

K. B. C. Brady, J. R. Shaulis and V. W. Skema 

Acid Mine Drainage: Field Applications 

Phosphatic Clay Slurries for Reducing Acid Mine Drainage From Reclaimed Mines Sites 44 

E. D. Chiado, J. J. Bowders and J. C. Sencindiver 

Laboratory and Field Testing of a Salt-Supplemented Clay Cap as an Impermeable Seal Over Pyritic Slates. 52 

E. W. Murray, S. P. Goudey, R. G. L. McCready, and J. Salley 

The Partitioning of Flow Components of Acidic Seeps From Surface Coal Mines and the Identification 

of Acid Producing Horizons within the Backfill 59 

D. T. Snyder and F. T. Caruccio 

The Use of Phosphate Materials as Ameliorants for Acid Mine Drainage 67 

J. J. Renton, A. H. Stiller and T. E. Rymer 

A Computer Simulation Probability Model for Geochemical Parameters Associated with Coal Mining 

Operations 76 

T. E. Rymer, II, J. J. Renton and A. H. Stiller 

Chemical Stability of Manganese and Other Metals in Acid Mine Drainage Sludge 83 

G. R. Watzlaf 

Acid Mine Drainage: General 

Treatments to Combat Pyrite Oxidation in Coal Mine Spoil 91 

C. A. Backes, I. D. Pulford and H. J. Duncan 

A Growth Inhibition Model for Thiobacillus Ferrooxidans 97 

K. L. Shuttleworth and R. F. Unz 

Rehabilitation of Waste Rock Dumps at the Rum Jungle Mine Site 104 

J. W. Bennett, J. R. Harries and A. I. M. Ritchie 

Chemical Inhibition of Iron-Oxidizing Bacteria in Waste Rock and Sulfide Tailings and Effect on Water 

Qual i ty 109 

G. R. Watzlaf 

Mapping Buried Tipple Refuse - Is the Magnetometer Better Than Terrain Conductivity? 117 

J. H. Schueck 

The Use of Pre-Aeration to Reduce the Cost of Neutralizing Acid Mine Drainage 131 

T. C. Jageman, R. A. Yokley and G. W. Heunisch 

Mine Geochemistry 

Methods for Determining Fundamental Chemical Differences Between Iron Disulfides From Different 

Geologic Provenances 136 

R. W. Hammack, R. W. Lai and R. J. Diehl 



Page 

Interpretation of Isotoplc Compositions of Dissolved Sulfates in Add Mine Drainage 147 

R. 0. van Everdingen and H. R. Krouse 

Geochemistry of Abandoned Lignite Mine Spoil in Texas 157 

J. K. Horbaczewski and F. Van Ryn 

The Rate of Oxidation of Pyrites From Coal and Ore Sources - An AC Impedance Study 164 

S. Chander and A. Briceno 

A Quantitative Assessment of the Natural Abatement of Acid Mine Drainage by Interaction 

with Allegheny Group Lithologies 170 

A. D. Stahl and R. R. Pan'zek 

Mine Waste Characterization 

Prediction of Add Mine Drainage From Duluth Complex Mining Wastes in Northeastern Minnesota 180 

K. Lapakko 

Acid Mine Drainage and Management of Eastern Oil Shale Resources and Associated Waste Products 191 

P. J. Sullivan and J. L. Yelton 

Evaluation of the TCLP Method for Two Mill Tailings 200 

J. C. Franklin and E. G. Zahl 

A Comparison of Results From Acid-Base Accounting Versus Potential Acidity Measured by the Peroxide 
Oxidation of Weathered and Unweathered Soils Containing Pyrite 206 

J. T. Ammons and P. A. Shelton 

The Interrelationship of Factors Influencing the Dissolution of Metals in Columns of Mine Tailings 210 

R. D. Doepker 

Improved Rock Durability Testing Techniques for Appalachian Valley Fills 220 

R. A. Welsh, Jr. and M. K. Robinson 

Mine Hydrology 

Acid Mine Drainage Research: Hydrology's Critical Role and Unifying Theme 228 

F. T. Caruccio 

Stream Sealing to Reduce Surface Water Infiltration into Underground Mines 232 

T. E. Ackman and J. R. Jones 

Unsaturated Fluid Flow in Mine Spoil: Investigative Methods Leading to a Quantitative Characterization. 240 

D. M. Diodato and R. R. Parizek 

Forecasting the Effect of Mine Site Rehabilitation Works on Local Ground Water Quality 248 

D. K. Gibson and G. Pantelis 

Photo-Linear Characterization, Lithologic Variability, and the Effects of Mining Activity by Fracture 
Studies and In Situ, Air-Injection, Permeability Testing 253 

C. A. Shuman and R. R. Parizek 

Modeling Sulfate Retention in a Lake Receiving Acid Mine Drainage 261 

A. T. Herlihy, A. L. Mills and W. Lung 

Pennsylvania's Approach to Cumulative Hydrologic Impact Assessment of Coal Mining Activities 269 

L. E. Langer 

Biological Mine Water Treatment 

Biological Metal Removal From Mine Drainage 274 

K. Nakamura 

Growth Responses and Iron Uptake in Sphagnum Plants and Their Relation to Acid Mine Drainage Treatment.. 279 

A. K. Spratt and R. K. Wieder 

An Overview of the Role of Algae in the Treatment of Acid Mine Drainage 286 

0. A. Kepler 

Trace Metal Removal From Stockpile Drainage by Peat 291 

K. Lapakko and P. Eger 



Page 

Nickel and Copper Removal From Mine Drainage by a Natural Wetland 301 

P. Eger and K. Lapakko 

Effects of a Sphagnum Peat on the Quality of a Synthetic Acidic Mine Drainage 310 

J. M. Dietz and R. F. Unz 

Wetland Systems for Mine Water Treatment: Case Studies 

Iron and Manganese Removal in a Typha -dominated Wetland During Ten Months Following Its Construction.... 317 

L. L. Stillings, J. J. Gryta and T. A. Ronning 

Constructed Wetlands for Acid Drainage Control in the Tennessee Valley 325 

G. A. Brodie, D. A. Hammer and 0. A. Tomljanovich 

The Siraco #4 Wetland: Biological Patterns and Performance of a Wetland Receiving Mine Drainage 332 

L. R. Stark, R. L. Kolbash, H. J. Webster, S. E. Stevens, Jr., K. A. Dionis, and E. R. Murphy 

Treatment of Discharge From a High Elevation Metal Mine in the Colorado Rockies Using 

an Existing Wetland 345 

J. C. Emerick, W. W. Huskie and D. J. Cooper 

The Tracy Wetlands: A Case Study of Two Passive Mine Drainage Treatment Systems in Montana 352 

M. T. Hi el and F. J. Ken'ns, Jr. 

Wetland Systems for Mine Water Treatment - Process & Design 

Effects of Cattails ( Typha ) on Metal Removal From Mine Drainage 359 

J. C. Sencindiver and D. K. Bhumbla 

Water and Soil Parameters Affecting Growth of Cattails: Pilot Studies in West Virginia Mines 369 

D. E. Samuel, J. C. Sencindiver and H. W. Rauch 

Determining the Capacity for Metal Retention in Man-made Wetlands Constructed for Treatment 

of Coal Mine Drainage 375 

R. K. Wieder 

Implications of Sul fate-Reduction and Pyrite Formation Processes For Water Quality in a Constructed 
Wetland: Preliminary Observation 382 

R. S. Hedin, P. M. Hyman and R. W. Hammack 

An Evaluation of Substrate Types in Constructed Wetlands Acid Drainage Treatment Systems 389 

G. A. Brodie, D. A. Hammer and D. A. Tomljanovich 

Poster Session Abstracts 

Isolation and Culture of a Manganese-Oxidizing Bacterium from a Man-made Cattail Wetland 399 

W. J. Vail, S. Wilson and R. K. Riley 

Screening of Mosses and Algae in Greenhouse Experiments for Their Ability to Remove Iron from Water 399 

H. J. Webster, L. R. Stark and S. E. Stevens, Jr. 

Ecological Engineering and Biological Polishing: Its Application in Close-out of Acid Generating 

Waste Management Areas 399 

M. Kalin 

Wetlands and Mine Drainage - An Ecotechnology Approach 400 

W. J. Mitsch, M. S. Fennessy, M. A. Cardamone, and D. Palmieri 

Peat Blanket to Lock Acid Mine Spoils in a Self-Sustaining Ecosystem 400 

A. Brown, S. p. Mathur and D. J. Kushner 

Wetland/Riparian Reconstruction Following Surface Mining: Part III. Recommendation on Channel 

Morphology to Support a Wet Meadow 400 

R. E. Behling 

A Staged Wetland Treatment System for Mine Water with Low pH and High Metal Concentrations 401 

T. M. Demko and B. G. Pesavento 

Construction of a Wetland Demonstration Site for a Metal-Mine Drainage 401 

T. R. Wildeman, L. S. Laudon and E. A. Howard 



Page 

Response of Thiobacillus ferrooxidans to Organic Compounds: Leakage of Cellular Material 

and Inh1 bltlon of Growth 401 

M. Bhatnagar and G. Singh 

Iron Monosulfide and Pyn'te Formation In Sediments of Lakes that Receive Acid Mine Drainage 402 

C. M. Wicks, J. S. Herman, A. L. Mills, and J. P. Schubert 

Hydrology and Geochemistry of Surface Coal Mine Lakes 402 

J. P. Schubert 

Water Resource Development Engineering and Acid Mine Drainage in the Upper Ohio River Basin 402 

M. Koryak 

Aeration Efficiency in the In-Line Aeration and Neutralization System 403 

T. E. Ackman and C. . C. Hustwit 

Treatment of Acidic Coal Refuse by Addition of Crushed-Stone Pond Screenings 403 

S. J. Stokowski, Jr. and R. R. Gilbert 

X-ray Diffraction Evaluation of Coal Overburden Neutralization Potential 403 

T. L. Despard 

Stream Sealing Over an Active Longwall Panel 404 

T. E. Ackman and C. C. Hustwit 

Micromap — A Database Retrieval and Display Manager for Microcomputers 404 

T. E. Rymer, II, J. J. Renton and A. H. Stiller 

Installation and Stability of Inverted Pyramid-Shaped Plugs for Closing Abandoned Mine Shafts 

Galena, KS Demonstration Project 405 

J. S. Volosin 



MINE DRAINAGE AND SURFACE MINE RECLAMATION 

Proceedings of a Conference jointly sponsored by the Bureau of 
Mines, the Office of Surface Mining Reclamation and 
Enforcement, and the American Society of Surface Mining and 
Reclamation. 



ABSTRACT 

Mine waste and mine reclamation are topics of major 
interest to the mining industry, the government and the general 
public. This publication and its companion volume are the 
proceedings of a conference held in Pittsburgh, April 19-21, 
1988. There were nine sessions (50 papers) that dealt with the 
geochemistry, hydrology and problems of mine waste and mine 
water, especially acid mine drainage. These comprise Volume 1. 
The nine sessions (43 papers) that dealt with reclamation and 
restoration of disturbed lands, as well as related policy 
issues, are included in volume 2. Volume 2 also contains the 
ten papers that pertained to control of subsidence and mine 
fires at abandoned mines. Poster session presentations are, in 
general, represented by abstracts; these have been placed in the 
back of both volumes. 



Reference to companies and specific products in these papers 
does not imply endorsement by the Bureau of Mines. 



USE of acid-base accounts in premining prediction of acid 



DRAINAGE POTENTIAL: A NEW APPROACH FROM NORTHERN WEST VIRGINIA 

o 

Richard S. diPretoro and Henry W. Rauch 



Abs trac t. --Premining prediction of postmining drainage 
quality is required by law, but the method most often used, in- 
volving acid-base accounting, was never intended to predict mine 
drainage quality and as used in the past has proven unreliable. 
In this study, postmining drainage chemistry was measured at 
base flow in a one-time sampling program for 75 surface coal 
mines in Monongalia and Preston Counties in West Virginia. The 
drainage chemistry was then compared to available acid-base 
account parameters to identify reliable premining predictors, if 
any, of drainage quality. Sampling locations included seeps, 
springs, first order streams, and inflows to ponds if they 
primarily contained mine drainage. Each sampling location was 
paired with the closest available acid-base account log, which 
was mathematically adjusted to yield the volume percentage of 
each geochemical unit in the mined coal overburden, plus one 
foot of pavement. Both total and net neutralization potentials 
(NP) of overburden are significantly and directly related to net 
alkalinity (alkalinity minus acidity) of mine drainage. Every 
sampled mine is producing drainage with positive net alkalinity 
if it has greater than 40 tons of CaCO^ equivalent of total NP 
per thousand tons of overburden, or greater than 30 tons of 
CaCOo equivalent of net NP per thousand tons of overburden. 
Conversely, every mine with less than 20 tons per thousand tons 
of total NP or less than 10 tons per thousand tons of net NP is 
producing drainage with negative net alkalinity. Total sulfur 
content and hence maximum potential acidity (MPA) in overburden 
bears very little relation to mine drainage quality. All mines 
in this study with very low (<0.2%) sulfur content in overburden 
produce acid drainage. The ratio of NP to MPA is significantly 
and directly related to net alkalinity of mine drainage. Every 
sampled mine is producing drainage with negative net alkalinity 
if it has an NP to MPA rati'o of less than 2.4. 



INTRODUCTION 

Public Law 95~87 (Surface Mining Con- 
trol and Reclamation Act 1977) requires 
premining planning. The law states that 
surface coal mining permits are not to be 
issued where required reclamation is not 
feasible or cannot be accomplished by the 
plan submitted. Acceptable drainage quali- 



Paper presented at the 1988 Annual 
Mine Drainage and Surface Mine Reclamation 
Conference sponsored by the American So- 
ciety for Surface Mining and Reclamation 
and the U. S. Department of the Interior 
(Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), April 
17-22 1988, Pittsburgh, PA. 

Richard S. diPretoro is a Consulting 
Geologist in Morgantown, WV and Henry W. 
Rauch is a Professor of Geology at West 
Virginia University, Morgantown, WV. 



ty is one of the components of required 
reclamation and such quality is to be in- 
sured in part by pre-iden t i f y ing acid- 
producing materials so as to avoid acid 
drainage. Thus, the Federal government has 
produced a legal requirement that acid- 
producing materials in coal overburden and 
pavement be identified in advance of mi- 
ning. To aid operators and state agencies 
in ^re-identifying acid-producing ma- 
terials, the U.S. Environmental Protection 
Agency (EPA) published a method developed 
at West Virginia University known as the 
acid-base account method (Sobek et al. 
1978). The method was intended to aid 
revegetation by identifying the best 
choices for topsoil substitute material. 
Although it was not intended by its authors 
to be used to predict postmining drainage 
quality, acid-base accounting has been used 
for that purpose and has produced mixed 
results (Renton 1985). 



This study attempted to identify re- 
liable premining predictors of acid 
drainage potential. This paper reports on 
reliability of acid-base account parameters 
for the premining prediction of acid post- 
mining drainage. 



Description of the Study Area 

The area of study for this project 
comprises all or part of the areas shown on 
thirteen 7.5 _ minute topographic quadrangles 
of the U.S. Geological Survey in northern 
Preston and eastern Monongalia Counties, WV 
(fig. 1). 



The coal-bearing rocks are middle 
Pennsylvanian to Permian in age and consist 
primarily of sandstone and shale with les- 
ser amounts of limestone and coal. The 
structures are low-lying open folds tren- 
ding northeast-southwest. 

The climate of the study area is of 
the humid, continental type characterized 
by mild summers and cold winters. Rainfall 
ranges from 42 to 52 inches (Herb et al. 
1981). All ground water that is eventually 
discharged from the study area is generated 
by the infiltration of precipitation except 
for minor amounts being flushed from deep 
saline aquifers. 




U. S. Geological Survey 
7.5 Minute Quadrangles 



1 Osage 

2 Morgantown North 

3 Lake Lynn 

4 Bruceton Mills 

5 Brandonville 

6 Rivesville 

7 Morgantown South 



8 Ma son town 

9 Valley Point 

10 Cuzzart 

11 Gladesville 

12 Newburg 

13 Kingwood 




Location of 
Study Area 



Figure 1. --Location map of the study area. 



METHODS OF INVESTIGATION 

Sampl ing 

Seventy-five surface water sites were 
selected for a one-time sampling program 
designed to indicate in a broad sense the 
quality of drainage from reclaimed surface 
coal mine sites. The mine sites were selec- 
ted based on pre-de t erm ined criteria dis- 
cussed below after examining over 400 per- 
mit applications for surface coal mines and 
other water and overburden data at the U.S. 
Office of Surface Mining Reclamation and 
Enforcement (OSM) in Morgantown, WVa. All 
reclaimed surface coal mines in Monongalia 
and Preston Counties in the Waynesburg, 
Upper and Lower Freeport, and Upper, 
Middle, and Lower Kittanning Coal Seams 
were selected for sampling where the mines 
couid be reached on the ground and had 
observable drainage. Seventy-five such 
sites were sampled. 

Mine drainage sampling locations in- 
cluded seeps and springs located in and 
near the mine backfill, both undeveloped 
and developed for domestic or livestock 
uses. First order streams and inflows to 
ponds were also sampled if they primarily 
contained mine drainage. The selected 
water sampling sites were usually near 
vegetated mine backfills that usually had 
been reclaimed for 2 or more years. They 
were mostly chosen on the downdip side from 
the mine drainage source with the highest 
flow rate. Where possible, they were estab- 
lished at or below the original level of 
the coal near the center of the mine to 
increase the chances of discerning any 
mining effects and to decrease the chances 
of dilution by unaffected ground water or 
of complicating effects from other nearby 
mines. Previously mined areas (before 
recent mining) were generally avoided, as 
were mines which had disposed tipple refuse 
or treatment residues. 

The field sites were sampled during 
base flow (at least 3 days after the last 
measurable rain), in the spring of 1985. 
Base flow is not subject to wide fluctua- 
tions and is indicative of shallow aquifer 
ground-water characteristics (Hem 1970, 
Todd 1980, Helsel 1983). The sampling 
sites were all located in heavily dissected 
upland areas not subject to any regional 
ground-water discharge. Temperature, spe- 
cific conductance, Eh, pH, and flow rate 
v/ere measured. Three water samples were 
collected and treated (as appropriate) for 
laboratory analysis at each site. 

Hot total acidity (HTA), alkalinity, 
sulfate, total and dissolved iron, and dis- 
solved manganese were determined (for each 
set of water samples) in the Hydrogeology 
Laboratory in the Department of Geology and 
Geography at West Virginia University. 

After the drainage sites had been 
sampled, acid-base accounts from OSM permit 
files were chosen to be paired where possi- 
ble with each water sampling point for 
later comparison. Such logs were taken 



from the same mine or a mine working the 
same coal seam within 2 km of the watershed 
draining to the sampling point. In most 
cases, logs fell within 1 km of the sam- 
pling point watershed as per recommenda- 
tions of the Suggested Guidelines (Surface 
Mine Drainage Task Force 1979). Such pairs 
were created for 30 out of the 75 mines 
s tudied . 



Adjustment of Core Log Data for Volume 

Core logs record the thickness and 
type of each rock lithology over an area 
that is so small compared to the area to be 
mined that they can be considered lines in 
cross-section. Mined overburden in the 
field area is not rectangular in cross- 
section, as it is in Illinois for example, 
but is more nearly triangular in cross- 
section because of the notable terrain 
slopes. Therefore, the volumes of each 
lithology represented by the core log are 
not in fact in the same volume proportion 
to each other in the mined overburden as 
they are in the linear record of the log. 
Those near the top are over-represented and 
those near the bottom are under-represented 
by an unadjusted log. 

Volumetric adjustment is in order if 
better predictions are to be made of the 
combined effects that the different over- 
burden rocks will have on water after mine 
reclamation, assuming that water quality is 
related to the mass of certain constituents 
in the rocks. The ideal solution would be 
to have a closely spaced grid of cores that 
would be s tratigraphically correlated, al- 
lowing the strata to be projected to their 
crop lines according to the dip and strike. 
The average area of each lithologic unit 
would then be measured by planimeter and 
multiplied by its average thickness to 
arrive at a total volume for each rock 
lithology. Tonnages would be computed if 
necessary, based on estimated (or measured) 
rock densities. 

An effort similar to that just des- 
cribed was successfully carried out and 
described in the literature by Caruccio and 
Geidel (1981). Such efforts were also 
carried out for some West Virginia surface 
mine reclamation permit applications for a 
brief time in 1981 (diPretoro 1986). The 
reason for such calculations in permit 
applications was to compute the amount of 
limestone that is supposed to be needed to 
compensate for deficiencies identified by 
acid-base accounting. 

For this study, a simpler volumetric 
adjustment was designed that could be ap- 
plied quickly to limited core log data. 
This volumetric adjustment was conceived by 
the senior author and carried out by means 
of a computer program written by Mr. Paul 
Sutter, systems analyst for the West Vir- 
ginia University Department of Geology and 
Geography . 

The volumetric adjustment yields the 
approximate percentage of the total over- 



burden volume occupied by each rock unit 
and requires knowing only the total 
thickness of the s tra t igraphic overburden 
column, the thickness of the individual 
rock units, and their relative positions. 
It starts with the recognition that every 
mined tract in the field area is part of a 
slope, hill, or mountain. It assumes that 
the shape of a mined hill can be approxi- 
mated as one of a number of three-dimen- 
sional shapes, each of which is a triangle 
in cross-section. These include shapes 
similar to those for contour mines and area 
or mountaintop removal mining operations 
where the top of the mountain is cone or 
prism shaped (fig. 2). The core log's top 
is usually located at the highest point of 
the triangle, and the coal or pavement 
forms the triangle's base leg which is 
roughly perpendicular to the core log. 
Symmetry about the core (vertical axis) is 
not required on conical shapes, and trian- 
gular cross-sections need not be isosceles. 
(In this study, some logs represented more, 
or less, overburden than was present at the 
mine producing the drainage samples. Where 
the log represented more overburden, the 
extra amount was removed from the calcula- 
tions. Where the log represented less 




CONTOUR 
HILLSIDE 



STRAIGHT 
HILLSIDE 




MOUNTAINTOP OR KNOB 



Figure 2 .-- Idea 1 i ze d shapes of hills 
having triangular cross-sections when 
mined . 



overburden, the missing amount could not be 
evaluated . 

The program takes the thickness of 
each layer of rock and calculates the per- 
centage of the area of the whole cross- 
sectional triangle occupied by that layer. 
The areas of the triangle and its parts are 
proportional to the volumes of the cor- 
responding parts of the overburden as a 
whole (fig. 3). 

Because the coal is, for the most 
part, gone after reclamation, its contribu- 
tion is removed and the relative percen- 
tages of each remaining layer are recalcu- 
lated. (Bottom coal, partings and other 
wastes are often left on the pavement but 
in the interests of simplicity are ignored 
in this study.) The percentage for each 
layer is then multiplied as a decimal frac- 
tion by the acid-base NP or MPA for that 
layer to arrive at its contribution to the 
total for each parameter. The adjusted NP 
and MPA for each layer is then summed to 
produce an adjusted NP and MPA for the 
entire column. Totals of unadjusted NP 

and MPA are physically meaningless because 
layer thickness and volume have not been 
taken into account. 

This right triangular method incor- 
porates simplifying assumptions. The core 
does not always represent all the overbur- 
den. The dip is ignored, but the maximum 
dip is low in the study area, rarely ap- 
proaching 10 percent. Cross-sections are 
not exactly triangular, but are slightly 
elliptical. Lithologic units are usually 
laterally discontinuous and may thicken, 
thin or pinch out altogether away from the 
core hole. Despite its shortcomings, how- 
ever, the volumetric adjustment technique 
used in this study is an adequate first 
approximation and is an improvement over an 
assumption of a linear relationship between 
percentage of a certain lithologic unit in 
a core log and its volume in the mine 
backfill. 



Characterization of Post-Mining Drainage 

For the purposes of this study, the 
legal definition of acid mine drainage 
(AMD) was used as a basis for characteriza- 
tion of postmining drainage. AMD is legal- 
ly defined as water discharged from an 
active, inactive, or abandoned mine and 
from areas affected by surface mining with 
a pH of less than 6.0 in which total acidi- 
ty exceeds total alkalinity (Code of Fed- 
eral Regulations 1986). The legal defi- 
nition of acid mine drainage creates an 
operational parameter to characterize water 
quality, herein called "net alkalinity," 
which is calculated by subtracting the 
acidity from the alkalinity of a drainage 
water sample with both measured in the same 
units (mg/L as CaCOo). Net alkalinity can 
be positive (when alkalinity exceeds acidi- 
ty) or negative (when acidity exceeds alka- 
linity). In this study, most drainage 
samples had only acidity or alkalinity. In 
the samples that had both acidity and alka- 



P = \ of Overburden Occupied by 

Stratum = i^fl£^2 x100 
Area WXYZ 




figure 3. — Triangular cross-sections of idealized hill 
used in the volumetric adjustment program. 



showing variables 



linity, one outweighed the other by a fac- 
tor of at least 10 in most cases. 

Net alkalinity for the sampled drain- 
age sites was compared by use of graphical 
data plots and statistical analyses with 
values for the geochemical parameters eva- 
luated. All the raw data were reported by 
diPretoro (1936); diPretoro and Rauch 
(1987) reported other results of the same 
study having to do with lithologic 
parameters . 



RESULTS AND DISCUSSION 

In spite of their shortcomings, acid- 
base accounts were used in this study be- 
cause they are available in large quantity 
and they do give some useful information on 
the geochemical nature of the individual 
strata . 



Neutralization Potential (NP) and Net NP 

Figure 4 compares net alkalinity of 
mine drainage to the acid-base account 
parameter of net NP in overburden, with 
each plotted point representing a sampled 
mine site. The statistically significant 
regression line depicts the average rise of 
net alkalinity of drainage with increasing 
net NP. Figure 4 also shows that all sites 
having net NP of less than 10 tons of 
calcium carbonate equivalent per thousand 
tons of overburden had drainage with nega- 
tive net alkalinity, while all sites having 
net NP of greater than 30 tons per thousand 
had drainage with positive net alkalinity. 
Note that the acid-base account method 
defines potentially toxic material in indi- 
vidual strata as that having a net NP of 
less than -5 (Sobek et al. 1978). This 
study indicates that the threshold level of 
concern for net NP for the overburden as a 



whole must be set higher. A similar trend 
(with a slightly lower R value that is 
statistically significant at 0.05 alpha) 
was also observed for total NP. All sites 
with total NP greater than 40 tons per 
thousand tons had drainage with positive 
net alkalinity while all sites with less 
than 20 tons per thousand tons had drainage 
with negative net alkalinity. 



Maximum Potential Acidity (MPA) 

MPA, calculated by multiplying total 
sulfur content as a decimal fraction by 
31.25 (Sobek et al. 1978, p. 55), was found 
to have no statistically significant rela- 
tionship (at 0.05 alpha) to net alkalinity 
of mine drainage. As shown in figure 5, 
there appears to be a slight rise in 
average net alkalinity with rising sulfur 
content, and all sites with very low sulfur 
produce acid. This seeming anomaly relates 
to the NP/MPA ratio present. NP of over- 
burden rises as MPA rises, but NP rises at 
a faster rate. Very low MPA values are on 
average associated with proportionately 
less NP (and hence neutralizing capability) 
than are higher MPA values (above 5). The 
median value of NP/MPA ratio for sites in 
this study with MPA below 5 is 0.5. The 
median value of NP/MPA ratio for sites with 
MPA above 5 is 2.9. 



NP/MPA RATIO 

Caruccio et al. (1980) and Caruccio 
and Caruccio (1982) have found a strong 
relationship between drainage quality and 
carbonate to sulfur ratio. In this study, 
the ratio of NP to MPA (herein called 
RATIO) was assumed to approximate the ratio 
of carbonate to sulfur. The regression 
line showing the relationship of RATIO to 
net alkalinity in mine drainage had the 



400 



200 



NET 
ALKALINITY e 
(mg/1 as 
CaC0 3 ) 



1400 



zero net alkalinity 




regression line 



KEY: 

• = data point 
2,3 = multiple data points 

REGRESSION LINE DATA: 



net alkalinity = -232 + 6.93 x NET 
R = 0.423 
n = 30 
significant at 0.05 alpha 



5 25 SO 75 iee 12 

ACID-BASE NET NEUTRALIZATION POTENTIAL 
(tons CaCO-j equivalent per thousand tons of material) 

Figure 4. — Net alkalinity in mine drainage versus acid-base net neutra- 
lization potential in coal overburden. 



greatest degree of statistical significance 
of those for the acid-base account para- 
meters tested in this study. Figure 6 
shows that all of 15 sites having a RATIO 
of less than about 2.4 had drainage with 
negative net alkalinity, while 11 of 14 
sites having RATIOS greater than about 2.4 
had drainage with positive net alkalinity. 
Note that the presently used acid-base 
account method would place the critical 
RATIO boundary below 1.0 (or even below 
zero) because geologic units having net NP 
greater than -5 are considered not poten- 
tially toxic (Sobek et al. 1978, Surface 
Mine Drainage Task Force 1979, Erickson et 
al. 1985). For example, one acid-base 
account in this study has an NP of 0.3 and 
an MPA of 4.6, producing a RATIO of 0.1. 
Strict application of acid-base account 
guidelines would result in this overburden 
being considered by some persons not poten- 
tially toxic because 0.3 minus 4.6 is 
greater than -5.0. The drainage from the 
two sampling locations within 2 km of this 
acid-base account, however, is highly acid. 



CONCLUSIONS 

This investigation of reclaimed sur- 
face coal mine sites in the northern West 
Virginia study area has led to four conclu- 
sions about the study results and a fifth 
general conclusion about acid-base 
accounting : 

1) After adjusting acid-base account 
core logs for volume, all sites with acid- 
base neutralization potential (NP) greater 
than 40 tons per thousand tons of overbur- 
den equivalent calcium carbonate yield 
drainage having positive net alkalinity; 
all sites with NP less than 20 tons per 
thousand tons equivalent calcium carbonate 
yield drainage having negative net alkali- 
nity . 

2) Acid-base maximum potential acidi- 
ty (MPA, calculated by multiplying percen- 
tage sulfur by 31.25) in premining coal 
overburden bears very little relation to 
net alkalinity of post-mining drainage in 
the study area except that below about 0.2 
percent sulfur, reclaimed overburden is 
likely to produce acid drainage because of 
concomitant low neutralizing capacity. 



■too 



NET 

ALKALINITY 

(mg/1 as 

CaC0 3 ) 



-698 




. zero net alkalinity 



KEY: 

• = data point 
2,3 = multiple data points 

n = 30 



*— 5 MPA line 



ACID-BASE MAXIMUM POTENTIAL ACIDITY (MPA) 

(calculated by multiplying percentage of total sulfur by 31.25; 

expressed as tons CaC03 equivalent per thousand tons of material) 

Figure 5. — Net alkalinity in mine drainage versus acid-base maximum poten- 
tial acidity in coal overburden. 



3) All sites with acid-base net NP 
(NP-MPA) greater than 30 tons per thousand 
tons of equivalent calcium carbonate (20 
percent of sites in this study) yield 
drainage having positive net alkalinity; 
all sites with net NP less than 10 tons per 
thousand tons calcium carbonate equivalent 
(47 percent of sites in this study) yield 
drainage having negative net alkalinity. 
The original acid-base account application, 
which placed the threshold level of net NP 
for drainage having positive net alkalinity 
at -5 (same units), was obviously in error. 

4) Acid-base NP divided by MPA yields 
a variable called RATIO. Sites with RATIO 
values greater than 2.4 are most likely to 
produce drainage having positive net alka- 
linity; sites with RATIO less than 2.4 are 
most likely to produce drainage having 
negative net alkalinity. The original 
acid-base account application which placed 
the threshold level of RATIO for drainage 
having positive net alkalinity at or below 
1.0, was in error. 



5) Despite its limitations, acid-base 
accounting can be used more effectively 
than it is currently being used to produce 
accurate premining assessments of post- 
mining drainage quality. Because of the 
complexity of coal mining and reclamation 
operations, professional judgement (taking 
into account lithology and history of 
drainage quality in a given area, as well 
as acid-base account parameters) will al- 
ways be required to arrive at a final 
determination . 



LITERATURE CITED 

Caruccio, F. T. and Caruccio, G. G. 1982. 
Method of Determining Weathering Char- 
acteristics of Rock Formations in 
Earth Moving Operations. United 
States Patent 4,328,001. 



200 



NET 
ALKALINITY 
(mg/1 as e 
CaC0 3 ) 




REGRESSION LINE DATA: 

net alkalinity = -427 + 138 x RATIO 
R = 0.486 
n = 29 
significant at 0.01 alpha 



12 3 4 5 

RATIO (Neutralization Potential/Maximum Potential Acidity) 

Figure b. -- Net alkalinity in mine drainage versus acid-base account RATIO 
(NP/MPA) in coal overburden. 



Caruccio, F. T., G. Geidel, and A. Pelle- 
tier. 1980. The Assessment of a Stra- 
tum's Capability to Produce Acidic 
Drainage. In: Proceedings of Sympo- 
sium on Surface Mining Hydrology, 
Sedimen tology and Reclamation. pp. 
437-443. University of Kentucky, 
Lexington, KY. 

Caruccio, F. T. and G. Geidel. 1981. Esti- 
mating the Minimum Acid Load that Can 
Be Expected from a Coal Strip Mine. 
In: Proceedings of Symposium on Sur- 
face Mining Hydrology, Sedimentology 
and Reclamation. pp. 117-133. Uni- 
versity of Kentucky, Lexington, KY. 



1986. Chapter 
pp. Office of 



Code of Federal Regulations 
30, Section 701.5. 622 
the Federal Register, National Ar- 
chives and Records Administration, 
Washington, DC. 

diPretoro, R. S. 1986. Premining Prediction 
of Acid Drainage Potential for Surface 
Coal Mines in Northern West Virginia. 
217 pp. Unpublished M.S. Thesis, De- 
partment of Geology and Geography, 
West Virginia University, Morgantown, 
WV 



diPretoro, R. S. and Rauch, H. W. 1987. 
Premining Prediction of Acid Drainage 
Potential for Surface Coal Mines in 
Northern West Virginia. J_n: Pro- 

ceedings of Symposium on Surface Mi- 
ning Hydrology, Sedimentology and Rec- 
lamation, pp. 395-404. University of 
Kentucky, Lexington, KY. 

Erickson, P. M., R. W. Hammack, R. L. P. 
Kleinmann. 1985. Prediction of Acid 
Drainage Potential in Advance of Mi- 
ning. I_n: Control of Acid Mine 
Drainage, Proceedings of a Technology 
Transfer Seminar. p.3~ll. Bureau of 
Mines IC 9027. 

Kelsel, D. R. 1983. Mine Drainage and Rock 
Type Influences on Eastern Ohio Stream 
Water Quality. Water Resources Bulle- 
tin 19:881-887. 

Hem, J. D. 1970. Study and Interpretation 
of the Chemical Characteristics of 
Natural Waters. Second Edition, 363 
pp. Water Supply Paper 1473, U.S. Geo- 
logical Survey. 



Herb, W. J., L. C. Shaw, and D. E. Brown. 
1981. Hydrology of Area 5, Eastern 
Coal Province, Pennsylvania, Maryland 
and West Virginia. 92 pp. Water Re- 
sources Investigation 81-538, U.S. 
Geological Survey, Harrisburg, PA. 

Renton, J. J. 1985. Evaluation of the Acid 
Producing Potential of Toxic Rock 
Materials. Mountain State Geology, 
pp. 7-12. West Virginia Geological 
and Economic Survey, Morgantown, WV. 

Sobek, A. A., W. A. Schuller, J. R. Free- 
man, and R. M. Smith. 1978. Field 
and Laboratory Methods Applicable to 
Overburdens and Minesoils. 203 pp. 
U.S. Environmental Protection Agency 
(EPA-oOO/2-78-054), Cincinnati, OH. 



Surface Mine Drainage Task Force. 1979. 
Suggested Guidelines for Method of 
Operation in Surface Mining of Areas 
with Potentially Acid-Producing Ma- 
terials. I_n: Symposium on Surface 
Coal Mining and Reclamation, pp. 177- 
197. Louisville, KY. 

Surface Mining Control and Reclamation Act. 
1986. Public Law 95"87, 30 USC 1201 
et seq. 142 pp. An Unofficial OSMRE 
Compilation, U.S. Department of the 
Interior, Office of Surface Mining 
Reclamation and Enforcement. 



Todd, D. 
535 

NY. 



K. 1980. Groundwater Hydrology, 
pp. John Wiley and Sons, New York, 



10 



EVALUATION OF OVERBURDEN ANALYTICAL METHODS AS MEANS TO PREDICT 
POST-MINING COAL MINE DRAINAGE QUALITY 1 



Patricia M. Erickson and Robert S. Hedin^ 



Abstract. --Prediction of the potential for acid drainage 
formation is required as part of the coal mine permitting 
process. Overburden analyses, by methods such as the acid-base 
account, are in common usage as part of this prediction despite 
the fact that a quantitative correlation between overburden 
properties and post-mining water quality had never been 
demonstrated adequately in the field. This study was designed to 
test the accuracy of predictions based solely on overburden 
analytical data. The study sites were selected to represent 
cases that have proven difficult to predict. Overburden samples 
were obtained from freshly exposed surfaces on highwalls or from 
cores preserved prior to mining. Post-mining drainage was 
analyzed at a reclaimed section of each site, paired to match the 
sampled overburden in terms of the presence and thickness of 
lithologic units. Overburden samples were analyzed by the acid- 
base account and a simulated weathering method. Overburden data 
were aggregated on thickness- and volume-weighted bases and water 
quality data were aggregated on a flow-weighted basis for each 
site. Drainage net alkalinity was correlated significantly 
(p<0.05) with the volume-weighted acid-base account parameters 
acid potential, neutralization potential, and net neutralization 
potential (r=0.358, 0.4197, 0.3799, respectively). However, 
when five sites having net neutralization potentials greater 
than 20 tons/1000 tons were excluded, these relationships became 
insignificant. Acid base account data showed boundaries useful 
in predicting the acid or alkaline character of post-mining 
drainage but no quantitative predictive equation could be 
developed. Simulated weathering test data correlated with 
actual sulfate concentrations in the post-mining seeps 
(r=0.4089, p=0.046), but the alkaline or acid nature of the 
seeps was not correlated with laboratory data (r=0.088, 
p=0.364). 



INTRODUCTION 



Ipaper presented at the annual meeting of 

the American Society for Surface Mining and Overburden analysis came into widespread use 

Reclamat 10 n, Apnl 17-22, 1988, Pittsburgh, PA. wUh pas$age of the ^^ Mining ^^ ^e 

2p a tricia M Erickson is a Suoervisory Reclamation Act of 1977 (Public Law 95-87). The 

„ . P ? c "VK.JrJP Inu^nEmLnJfi ] aw requires, as part of the permitting process, 

Physical Scientist with the Environmental the identificatio f; of potentially acid producing 

Technology Group of the U.S. bureau of Mines, strata and determinatio P n of the J robabi ; nydrol ^ gic 

Pittsburgh Research Center Dr. Robert S. Hedin consequences of mining, including water resource 

is an Ecologist working with the Environmental degradation. Because little research had been 

Technology Group, Pittsburgh Research Center conducted on water quality prediction, the 

under a post-doctoral fellowship with Oak Ridge. regulatory authorities and industry initially had 

to adapt available methods to these purposes. 



11 



The acid-base account method of overburden 
analysis was developed in the early 1970' s for 
identification of strata that could be used as soil 
substitutes in revegetation of coal mines (Smith et 
al 1974). The method (described subsequently in 
this paper) is based on determining whether the 
capacity of a stratum to produce acid via pyrite 
oxidation is greater or less than the ability of 
the same stratum to neutralize strong acid. A 
neutral izer deficiency greater than 5 tons/1000 
tons as calcium carbonate equivalent^, a figure 
common for natural soils in West Virginia, was the 
recommended standard for identifying strata 
unsuited for supporting vegetation. 

This method was adopted widely as the 
overburden analytical tool for identification of 
acid and toxic strata. The term "toxic" was meant 
to indicate toxic to plants by virtue of high 
acidity levels. While the purpose and application 
of the method was delineated clearly in the 
literature, its more recent use as an indicator of 
water quality is not clearly defined. 



Based on the 
acid-base account 
and predict wheth 
would comply with 
other overburden 
numerous alternat 
static or kinetic 
properties while 
rates of relevant 
tests are kinetic 
the laboratory th 
field setting. C 
have recently bee 
Erickson, 1987) 



real or perceived failure of the 
to identify acid-producing strata 
er or not post-mining discharges 
regulatory standards, interest in 
analysis methods has grown. The 
ive methods can be classified as 
; the former rely on bulk rock 
the latter take into account the 
reactions. Simulated weathering 
methods that aim to accelerate in 
e reactions that take place in the 
ommon static and kinetic tests 
n reviewed (Ferguson and 



At the start of this study, none of these 
methods had been adequately evaluated by comparing 
the predicted water chemistry to actual drainage 
chemistry. That is, there was no procedure for 
interpreting the results to predict post-mining 
water quality or character with a known degree of 
accuracy. The purpose of this study was to define 
the predictive capability of overburden analytical 
methods by applying them to many mine sites. Post- 
mining water quality measurements were made at the 
discharge points from reclaimed mine sections. 
Pre-mim'ng overburden samples were collected from 
preserved cores or from fresh channels cut in 
extant highwalls. These samples were analyzed 
using the acid-base account method and a simulated 
weathering test. The results of both overburden 
tests were then compared with the water quality 
results to determine if significant correlations 
exist. 



^The unit, tons/1000 tons, derives from 
agronomy where 1000 tons is the estimated weight of 
an acre-plow layer of soil. The unit is identical 
to parts per thousand. CaC.03 equivalent indicates 
the mass of calcium carbonate needed to yield a 
specific magnitude of electrical charge (CaC03 = 
50 g/mol of charge). The unit is used so that 
terms of opposite charge, such as acidity and 
alkalinity, are additive. 



METHODS 

Technical Approach 

Thirty-two sites were selected for the study, 
each comprised of a pre-mining section and a post- 
mining reclaimed section, so that pre- and post- 
mining data could be gathered simultaneously. The 
general concept was developed at the Bureau of 
Mines while the detailed approach was developed by 
Engineers International in response to our request 
for proposals. The contractor and its 
subcontractor, Sturm Environmental Services, 
selected all sites, conducted all sample collection 
and performed all water and rock analyses. Bureau 
personnel conducted all data analysis following 
completion of the contract. 

Sites were selected to represent cases that 
are difficult to predict. Overburden columns 
completely dominated by either acidic or alkaline 
strata were eliminated from consideration. Sites 
that were backfilled or reclaimed using acid- 
mitigating methods, such as selective spoil 
placement or surface alkaline addition, were also 
excluded. To minimize effects of local geology, 
sites in Pennsylvania, West Virginia, Maryland, 
Illinois and Kentucky were included in the study. 
Sites in west-central and southwestern Pennsylvania 
comprised about half of the study sites, through a 
prevalence of suitable candidates rather than 
through design. 



Water Quality 

Sampling and Analysis 

Post-mining water quality data were collected 
from reclaimed spoil sections at 32 mine sites 
paired with overburden sampling sites. Water 
sampling and flow measurement were conducted at all 
discharge points identified during the initial 
visit to each site. The discharges that comprised 
the majority of the flow, usually two or three at 
each site, were sampled subsequently at least nine 
times over about a year. Primary measurement and 
analysis methods for parameters used in this report 
are shown in Table 1. 



Table 1. --Analytical methods for selected 
parameters (U.S. EPA 1983) 

Parameter Measurement Method 

Water flow Bucket and stopwatch, L/min 

pH Electrometric, standard units 

Acidity Titration (hot, peroxide) to pH 

8.3, mg/L as CaC03 
Alkalinity Titration (cold) to pH 4.5, 

mg/L as CaC.03 

Net Alkalinity Alkalinity minus Acidity, mg/L 

Sulfate Col orimetric methyl thymol blue, 

mg/L 



12 



Data Aggregation 

Water quality data for each site were 
aggregated to produce a single value for each 
parameter by calculating a flow-weighted average: 



flow-weighted X = 
n 
Y (concentration of X)(flow) 



(1) 






(flow) 



where n is the total number of samples at the site 
for which flow and concentration of X were measured 
and X is the chemical quantity net alkalinity, 
sulfate, or iron in mg/L. 



Overburden 



Sample Collection 



Overburden samples were collected in two ways 
to represent pre-mining conditions. At nine sites, 
well-preserved and well -documented core samples 
taken prior to mining of the reclaimed section of 
the site were used for analysis. In all other 
cases, a highwall that could be determined to match 
the backfill, in terms of the presence, thickness, 
and appearance of each lithologic unit, was sampled 
for overburden analysis. Highwall samples were 
collected from a channel of freshly exposed 
surfaces, with lateral displacement at lithologic 
boundaries where needed to represent the stratum as 
records showed it had existed in the reclaimed 
section prior to mining. The entire length of each 
lithologic unit was sampled to account for intra- 
stratum differences in composition. An average of 
eight rock units (individual strata or adjacent 
strata of similar appearance) were collected from 
each site. 

Acid-base Account 

Acid-base account analyses were conducted 
according to procedures developed at West Virginia 
University (Smith et al 1974, Sobek et al 1978) 
with minor modifications. The portion of the 
procedure that produced data used in this paper is 
as follows. Samples were air-dried, mixed, crushed 
to pass a 60-mesh sieve, and stored in sealed 
containers until analyzed. 

Acid potential (AP) is defined as the amount 
of acid that would be produced if all pyrite in the 
sample reacted according to the following 
stoichiometry: 

FeS 2 + ^2 + 2 H 2°' = Fe ( 0H )3(s) + 2S0 4 + 4H+ ( 2 ) 

One mole of pyrite, containing 64 g sulfur, 
produces 4 moles of H + acidity, equivalent to 200 g 
CaCOj acidity. Thus, the acid potential of pyritic 
sulfur is 3.125 g acidity (as CaCO^) per g sulfur; 
1 pet pyritic sulfur has the potential to produce 
31.25 tons/1000 tons acidity. Total sulfur 
concentrations were measured with an automated 
Fisher sulfur analyzer, consisting of a furnace and 
SO2 titrator, according to the instrument 
instructions. Sulfur speciation was analyzed for 
samples containing more than 0.5 pet sulfur by hot 
extraction with hydrochloric and nitric acid to 
remove sulfate and pyritic sulfur, respectively. 



Organic sulfur was defined as the sulfur fraction 
remaining after acid extraction. AP was calculated 
from both total sulfur and, when analyses were 
available, pyritic sulfur. Total sulfur is often 
used in mine permit applications to minimize 
analytical costs. 

Neutralization potential (NP), defined as the 
ability of the stratum to neutralize strong acid, 
was determined by treating a 2 g sample with 20 to 
80 mL of 0.1 M HC1 , heating nearly to boiling, and 
swirling periodically until no gas evolution was 
observed. The samples were made up to 125 mL with 
distilled water, boiled for 1 min, and cooled to 
room temperature. The treated sample was then 
titrated with standard NaOH (0.1 or 0.5 M) to pH 7. 
NP was calculated as the amount of HC1 consumed by 
the sample and converted to the units of tons 
CaC03/1000 tons material: 



NP, tons/1000 tons = 

g HC1 consumed 50 g CaC0 3 



g sample 



36 g HC1 



x 1000 



(3) 



Net neutralization potential (NETNP) was 
calculated for the stratum by subtracting AP from 
NP. A positive NETNP indicates an excess of 
neutral izers while a negative NETNP indicates a 
deficiency of neutral izers in the stratum. 

Simulated Weathering 

A simulated weathering procedure was applied 
to all samples from 18 of the 32 study sites. The 
subset of sites, necessary to save cost, was 
selected to cover the range of post-mining water 
quality observed at all sites and to include sites 
for which true water quality and acid-base account 
results conflicted. 

Samples were crushed to pass a 2-mm sieve and 
300 g portions were placed in a thin layer on the 
bottom of weathering chambers. The chambers were 
each rinsed with 300 mL of distilled water; the 
collected leachate was labelled "week 1" and 
analyzed. An inoculant of 10 mL of acid mine 
drainage was then added to each sample to establish 
a population of iron-oxidizing bacteria. For the 
next 5 to 8 weeks, leaching was performed on a 
7-day cycle: dry air was passed over the sample 
for 3 days; humidified air was passed over the 
sample for 3 days, and leaching occurred on the 
last day, by the addition of 300 mL of distilled 
water for 1 hr followed by draining. Any fines 
washed out during draining were collected by 
centrifugation, resuspended in a small amount of 
water, and returned to the chamber. The leachate 
volume was measured and portions were analyzed for 
pH, acidity, alkalinity, iron, and sulfate (table 
1). 

Leachate composition from each sample was 
aggregated over the duration of the weathering test 
to calculate cumulative loads of pollutants. 
Cumulative load of net alkalinity (alkalinity minus 
acidity) was calculated according to: 

Cumulative net alkalinity, mg = 

n 1 

T (net alkalinity, mg/L) (leachate volume, L)l (4) 



13 



where n is the total number of weeks of leaching. 
Cumulative sulfate load was calculated in the same 

manner. 

The same simulated weathering procedure was 
also applied to composite samples from 5 of the 
study sites. Composite samples were made by mixing 
all the strata from a site in weight ratios equal 
to the thickness ratios of the strata in the 
overburden column. Equation (4) was again used to 
calculate cumulative product loads. 

Data Aggregation 

Overburden analytical data from both the acid- 
base account and single-stratum simulated 
weathering tests were aggregated in two ways to 
produce a single value to represent the entire 
overburden column at each mine. First, a 
thickness-weighted average was calculated as 

thickness-weighted NETNP = 



n 



[(NETNP)(thickness)] 



(5) 



ii 

r" 



(thickness) 



where n is the total number of strata sampled at 
the site. 

Second, a volume-weighted average, modelled 
after diPretorio's (1986), was calculated by 
substituting an estimated volume for thickness in 
the previous calculation. The volume of each unit 
was estimated by assuming that the mine cross- 
section was an isoceles right triangle, having two 
sides equal to the height of the overburden column. 
Using geometry, these assumptions yield the cross- 
sectional area of a stratum: 

2 2 
Area = 1/2 (D 2 - Dj ) 



(6) 



where Dj and D 2 are the depths of the top and 
bottom of the stratum, respectively. 



RESULTS AND DISCUSSION 



Water Quality 



Descriptive statistics for selected water 
quality data are summarized at the individual 
sample and mine levels in Table 2. Samples 
collected from a single site often showed changes 
in water quality during the study period. Many 
sites had seeps of both acidic and alkaline 
drainage (fig. 1), and individual seeps 
occasionally changed character during the study 
period (fig. 2). Peak contaminant concentrations 
and loads occurred at different times of the year 
in different seeps (fig. 2). These observations 
are a reflection of the heterogeneity of both the 
spoil material and its local environment within the 
backfill (Ladwig and Campion; Lusardi and 
Erickson). They also illustrate the difficulty 
associated with using aggregated values to 
represent the entire mine site. However, as 
drainage from a site usually flowed into one water 
course, it seemed reasonable to attempt to weight 
the water quality parameters by flow to represent 
the site discharge. 

Mean and median values of net alkalinity, 
iron, and sulfate were not in accordance with the 
stoichiometry shown in equation (2). Equation (2) 
produces 25 parts net acidity (or 25 parts net 
alkalinity) per 24 parts sulfate and no dissolved 
iron. Different stoichiometric equations describe 
pyrite oxidation at the low pH values that were 
observed at several sites: 



FeS 2 + 14Fe 3+ + ^0 2 + 8H 2 = 



15Fe 2+ + 2S0 4 + 16H+ 



14 



14Fe 2+ + ^2 + 14H+ " 14Fe + 7H 2° 



FeS 



15 



2+ 



2 + 4°2 + H 2 = Fe^ + + 2S0 4 + 



2H + 



(8) 



(9) 



(10) 



Assuming constant length and constant density 
for all strata, the proportion of spoil produced 
from an individual overburden unit was calculated 
as: 



Fractional Volume 



2 2 
02 - Pi 

2 
DT 



(7) 



where Dj is the total depth to coal 



Data Analysis 

Data were entered and manipulated in Fortran- 
readable files on a VAX mainframe computer. 
Descriptive statistics, Pearson correlations and 
linear regression analyses were performed using 
SPSS-X software. The significance of correlations 
was judged at the 0.05 probability level. 



Equations 10 and 2 yield the same ratio of acidity 
to sulfate (2 moles H + acidity and 2 moles latent 
H + acidity in the Fe 2+ produced in equation 10 
versus 4 moles H + acidity in equation 2). Equation 
10 produces 25 parts net acidity and 7 parts iron 
per 24 parts sulfate. 

Neutralization of pyrite oxidation products by 
carbonate minerals (e.g., CaC03 + 2H + — > Ca 2+ + 
H 2 + C0 2 (g)) can decrease acidity relative to 
sulfate in the drainage. The accompanying rise in 
pH causes iron hydrolosis and precipitation. Only 
drainage from highly pyritic and neutral izer-free 
wastes, such as aged coal refuse, approaches the 
stoichiometry expected from equation (10). 



14 



Table 2. — Descriptive statistics for selected post-mining water quality variables. ^ 



Parameter 



Mean 



Median 



Range 



All individual samples (n = 797) 
Net alkalinity, mg/L as CaOC>3. 

Iron, mg/L 

Sulfate, mg/L 



Flow-weighted mine aggregate 
Net alkalinity, mg/L as CaOC>3. 

Iron, mg/L 

Sulfate, mg/L 



-109. 

14.1 
1090. 



-130. 

10.9 
1140. 



13.0 
2.68 
805. 



-30.6 

4.43 
826. 



-2620. - 620. 
0.00 - 99.0 
23.0 - 4350. 



-1630. - 536. 
0.034 - 50.1 
37.4 - 4020. 



-•-Values rounded to 3 significant figures. 





Figure l.--The major seeps at site 21 discharged 
both acid and alkaline water. 



Figure 2. --Seep alkalinity at site 4 varied 
laterally and temporally. 



Carbonate minerals in equilibrium with clean 
water at atmospheric pCO? are expected to yield 
less than 100 mg/L net alkalinity. Yet 
alkalinities measured at several sites in this 
study were higher, with a maximum of 620 mg/L 
(Table 2). Elevated pCC^, possibly produced by 
acid dissolution of limestone, increases the 
solubility of carbonate minerals. Elevated pCC>2 
levels have been measured in acidic coal mine 
spoil . 



Aggregated at the mine level, net alkalinity 
ranged from -1630 to +536 mg/L, while mean sulfate 
concentrations ranged from 37.4 to 4020 mg/L (table 
2). Seventeen of thirty-two sites produced acidic 
drainage. These values indicate that the site 
selection methods were adequate to identify mines 
that produced drainage of both acceptable and 
unacceptable quality. 



15 



Table 3. --Descriptive statistics for acid-base 
accounts. * ■ 



iDt 



Parameter 


Mean 


Median 


Ranqe 


All samples 

AP 2 


21.6 


5.84 


0.031 


- 258. 


NP 


37.6 


8.41 


0.22 


- 937. 


NETNP 


15.8 


1.77 


-257. 


- 936. 


Thickness-weighted 










mine aggregates 
AP 


11.8 


9.36 


1.04 


- 34.2 


NP 


38.1 


13.4 


3.43 


- 272. 


NETNP 


26.2 


5.81 


-22.7 


- 252. 


Volume-weighted 
mine aggregates^ 
AP- 










17.4 


14.3 


1.57 


- 51.8 


NP 


39.1 


15.3 


3.30 


- 197. 


NETNP 


21.6 


4.01 


-31.6 


- 157. 



*A11 values rounded to 3 significant figures or 

2 decimal places. 
2 Total sulfur values that 

detection limit were reco 

rise to a single-stratum 
^Site 13 was omitted since 

available to estimate vol 



registered bel 
)rded as 0.001 

AP value of 0. 
insufficient 
lumes. 



ow the 
pet, giving 
031. 
data were 



Table 4. --Correlation between flow-weighted water 
quality parameters and overburden variables 
derived from acid-base account data. 



Overburden 
variables, 

tons/1000 tons 
Thickness-weighted 

AP 

NP 

NETNP 

NP/AP 

Volume-weighted 

AP 

NP 

NETNP 

NP/AP 



Water Variables 

Correlation coefficients 

(probabil ity) 



Flow-weighted 

net alkal inity 

mg/L as CaC0 3 



0.3050 (0.045) 
0.2562 (0.078) 
0.2225 (0.110) 
0.0701 (0.352) 

0.3580 (0.024) 

0.4197 (0.009) 

0.3799 (0.018) 

0.2096 (0.129) 



Flow-weighted 

sulfate 
mg/L as SO 4 



0.1655 (0.183) 
0.2785 (0.061) 
0.2672 (0.070) 
0.2001 (0.136) 

0.0845 (0.326) 
0.3701 (0.020) 
0.3866 (0.016) 
0.2986 (0.051) 



Acid Base Account 

Net neutralization potential (NETNP, equal to 
the neutralization potential less the acid 
potential) was calculated for each overburden unit 
from each site using total sulfur data. 
Descriptive statistics are shown in Table 3. A 
wide range of acid-base properties were exhibited 
in the data set. Eighty-five of 264 individual 
strata were potential acid-producers according to 
the 5 ton/1000 ton neutralization deficit (NETNP<-5 
tons/1000 tons) used for revegetation. Every site 
had at least one rock unit that met this criterion 
for acid potential, yet only 17 of the sites 
produced acidic drainage. Therefore, the presence 
of a stratum having NETNP<-5 tons/1000 tons was not 
a useful means to predict acidic drainage. 

Post-mining alkal inity and sulfate 
concentrations were correlated with several acid- 
base account parameters (Table 4). The 
correlations generally were stronger when the acid 
base account values were aggregated by overburden 
volume rather than thickness. This is expected 
because the strata are represented more truly in 
the former aggregation even though the mine 
geometry assumptions were crude. Volume-weighted 
AP, NP, and NETNP were all significantly related to 
drainage net alkalinity, while only NP and NETNP 
were significantly related to drainage sulfate. 
Even when AP was calculated from pyritic sulfur 
values the correlation was not significant at the 
5 pet level (unpublished data from this study). 
There was only a marginally significant 
relationship between sulfate concentration in the 
drainage and the value of NP/AP (r=0.2986, 
p=0 . 051 ) , intended to represent an oxidation 
potential inhibited by neutralization potential. 
All these correlations became insignificant 
(p=0.073 to 0.397) when the five sites having NETNP 
higher than 80 tons/1000 tons were excluded from 
the analysis. Including the acid and alkaline end- 
members of the set of all possible overburden 
compositions is expected to yield a higher 
correlation coefficient because overburden columns 



totally dominated by acid or alkaline strata 

produce analogous drainage quality. However, these 

correlations fail in the crucial hard-to-predict 
cases. 

The significance of the correlations shown in 
Table 4 did not change markedly when AP and NETNP 
values were calculated using pyritic sulfur rather 
than total sulfur values. For example, the 
correlation coefficient for water alkalinity and 
thickness-weighted NETNP only increased from 0.2225 
to 0.2262 when pyritic sulfur values were used. 
This finding results from the fact that pyritic 
sulfur and total sulfur contents in these samples 
were highly correlated. If we had only considered 
those cases where organic sulfur content was high 
and represented a variable fraction of total 
sulfur, AP and NETNP values based on pyritic sulfur 
rather than total sulfur would be expected to yield 
significantly stronger correlations. 

Sulfate production in the reclaimed spoil may 
be affected by inherent differences in pyrite 
reactivity (Hammack 1988), availability of oxygen 
(Lusardi and Erickson), juxtapositioning of 
different rock units in the backfill and other 
factors. Splits of the samples from this study 
have been analyzed for thermal pyrite reactivity by 
evolved gas analysis and the data will be used to 
determine if an effective pyrite content (some 
combination of sulfur content and reactivity 
variables) can better predict drainage quality. 

A scatterplot of drainage alkalinity versus 
volume-weighted NETNP (fig. 3) illustrated the lack 
of predictive value in this overburden variable 
alone. However, boundaries can be drawn to 
delineate the acid or alkaline character of the 
drainage. At NETNP values greater than 80 
tons/1000 tons, all sites produced alkaline 
drainage. No such clear boundary was found on the 
acid end of the spectrum, probably because the data 
set did not include equally acid-dominated sites. 
Such sites usually fail to gain permit approval. 



16 






\ 

0) 

E 



< co 

<C0 
co 

h- = 

LU ° 

UJ 



< 

UJ 











D 


0.4 - 




a 


D 




0.2 - 






a D 




n n 


D 

a 


° a^n D 


a 




U.U 


a odd 


□ D 






-0.2 - 




D 






-0.4 - 


D 

a 


Si 
□ 






-0.6 - 










-0.8 - 


c 


] 






-1.0 - 










-1.2 - 




a 






-1.4 - 










-1.6 - 

1 A 




D 






It) 


i i i 


i i i i i i i 


i i i i i i i 


i 



-40 



-20 



20 40 60 80 100 

NET NEUTRALIZATION POTENTIAL. t/IOOOt 



120 



140 



160 



Figure 3. — Scatterplot of volume-weighted net alkalinity versus volume-weighted NEINP for all sites. 



Intermediate boundaries are arbitrary: 



0<NETNP<20 
NETNP<10 
NETNP< 



acidic drainage in 8 of 15 sites 
acidic drainage in 15 of 20 sites 
acidic drainage in 9 of 11 sites 



Clearly, sites having volume weighted NETNP values 
lower than 20 tons/1000 tons were dominantly acid 
producers. 

Simulated Weathering 

The simulated weathering procedure was 
applied to all overburden samples from 18 of the 
study sites. Results of weathering tests, along 
with acid-base account results for the same sites 
are shown in Table 5. 

Correlations were calculated that related 
weathering test parameters to actual drainage 
parameters (Table 6). The only significant 
relationship was between weathering sulfate and 
drainage sulfate (r=0.4059, p=0.046). These 
correlations suggest that the laboratory procedure 
somewhat simulated sulfate production as intended 
but failed to simulate subsequent neutralization 
processed that increase alkalinity and precipitate 
iron. As with the acid-base account, the direct 
correlation was not strong enough to predict post- 
mining drainage quality. 



Failure to mimic the neutral i 
and thereby predict drainage alkal 
concentrations, could result from 
tests being conducted on individua 
atmospheric pCC>2 levels. Both ele 
sequential positioning of alkaline 
strata affect the amount of carbon 
dissolved. No leachate sample exc 
net alkalinity, while many post-mi 
samples had net alkalinity in the 
620 mg/L. 



zation processes, 
inity and iron 
the laboratory 
1 strata at 
vated CO? and 
and acidic 
ate minerals 
eeded 100 mg/L 
ning drainage 
range of 100 to 



Samples from five of the study sites were also 
weathered in the laboratory as thickness-weighted 
composites of all overburden strata. Leachates 
from these combined samples sometimes exceeded 100 
mg/L net alkalinity. While conclusions drawn from 
such a small data set are tenuous, this result 
suggests the intriguing possibility that 
neutralization reactions that occur in the field 
can be simulated in the laboratory if the 
neutralizing strata are exposed to leachate from 
acidic strata. 



17 



Table 5. — Summary of data for sites analyzed by the simulated weathering test 1 









Pre-minina; Data 




Simulated weathering 


Acid-base account 




Post-mining Water Quality 


volume-weighted 


volume-weighted 


Site 


flow-weiqhted, mq/L 


cumulative load, mq 


tons/1000 tons 


ID 


Alkalinity 


Sulfate 


Alkalinity 


Sulfate 


AP 


NETNP 


32 


-1150. 


4020. 


-40.2 


239. 


5.54 


18.8 


14 


-903. 


1010. 


7.66 


48.3 


7.31 


-0.49 


29 


-509. 


1430. 


-41.6 


106. 


11.6 


1.92 


18 


-453. 


695. 


-210. 


208. 


19.2 


-10.4 


12 


-278. 


1940. 


2.08 


37.5 


2.35 


2.99 


19 


-176. 


819. 


24.0 


67.7 


1.57 


1.73 


8 


-109. 


737. 


-17.9 


146. 


13.0 


-6.75 


28 


-81.3 


741. 


-4.69 


57.5 


6.75 


8.18 


4 


-72.6 


832. 


17.6 


54.8 


15.0 


4.38 


21 


-36.8 


761. 


26.5 


108. 


14.8 


-3.54 


31 


64.8 


258. 


24.8 


83.7 


18.3 


1.86 


11 


87.4 


332. 


-174. 


281. 


32.5 


-16.8 


6 


91.7 


797. 


39.5 


72.2 


13.5 


17.8 


13 2 


109. 


259. 


22.7 


27.1 


1.47 


7.25 


5 


170. 


1040. 


17.7 


92.7 


22.0 


83.2 


22 


201. 


2280. 


-280. 


995. 


51.8 


145. 


25 


313. 


695. 


20.7 


199. 


29.9 


11.2 


23 


536. 


3310. 


41.6 


315. 


25.8 


157. 



^All values rounded to 3 significant figures or a maximum of 2 decimal places. 
2 Site 13 is thickness-weighted rather than volume-weighted. 



Table 6. --Correlations between flow-weighted water quality parameters and 
simulated weathering test parameters. 



Overburden variables, 


Correlation coefficient (probability) 
Water variables, flow-weiqhted mq/L 


cumulative load, mq 


Net alkalinity 


Sulfate 


Iron 


Straight average 

Net alkal inity 

Sul fate 


-0.0436 (0.432) 
0.2554 (0.153) 

0.0880 (0.364) 
0.2234 (0.186) 


0.1633 (0.259) 
0.2308 (0.178) 

-0.0861 (0.367) 
0.4089 (0.046) 


0.0193 (0.470) 
-0.1680 (0.253) 

-0.1033 (0.342) 
-0.2276 (0.182) 


Volume-weighted 

Net alkal inity 

Sulfate 



18 



CONCLUSIONS 



LITERATURE CITED 



The results of th 
overburden analysis me 
adequately predict pos 
when used alone. This 
overburden analytical 
the pre-mining site as 
regulatory authorities 
experience in judging 
of a proposed mine. 



is study confirmed that the 
thods tested do not 
t-mining drainage quality 

result is not surprising, as 
data is only one component of 
sessment. Mine operators and 

utilize other data and past 
the environmental suitability 



Examinatio 
a high degree o 
in contaminatio 
the data showed 
accompanied by 
While these fin 
literature and 
us that predict 
the whole mine 
seep water qua! 



n of water quality data alone showed 
f temporal and lateral variability 
n levels at a single site. Further, 
that pyrite oxidation was 
varying degrees of neutralization, 
dings are consistent with the 
fully expected, they serve to remind 
ion of overall water quality from 
system and prediction of individual 
ity are two different things. 



Results obtained from the acid-base account 
confirmed that the use of a 5 ton/1000 ton 
deficiency of neutral izers (NETNP < - 5 tons/1000 
tons) to delineate acid-producing strata is invalid 
as a predictive tool. Each of the sites used in 
this study contained at least one lithologic unit 
that was potentially acid-producing by this 
standard. Yet only about half the sites produced 
acidic drainage. 

Qualitative boundaries were found for the 
acid-base account to delineate the alkaline or 
acidic character of the overburden. All sites 
having a volume-weighted NETNP greater than 80 
tons/1000 tons yielded alkaline drainage. In 
contrast, the majority of sites having a volume- 
weighted NETNP less than 20 tons/1000 tons yielded 
acid drainage. Unfortunately, no sites were found 
having NETNP in the range of 20 to 80 tons/1000 
tons; this range is left as a gray area in 
predicting acidic or alkaline drainage. Additional 
data, already available in mining company and 
regulatory agency files, should be used to test and 
further refine these boundaries. 

The weathering procedure used in this study 
failed to yield an accurate prediction of water 
quality and failed to delineate boundaries for 
drainage of differing acid/alkaline character. The 
correlation between sulfate produced during 
simulated weathering and sulfate concentrations 
observed in the field suggested that pyrite 
oxidation can be mimicked in the laboratory. 
Alkalinity-generating processes were not mimicked 
by this method. Limited data from laboratory 
weathering showed at least a potential for better 
simulation of neutralization processes when the 
overburden strata were composited into a single 
sample prior to weathering. 

It is important to pursue improved means of 
predicting post-mining drainage quality because of 
the environmental and economic repercussions of 
incorrect predictions. 



diPretoro, R. S. 1986. Premining prediction of 
acid drainage potential for surface coal mines 
in northern West Virginia. Master's thesis, 
West Virginia University, Morgantown, WV, 
217 p. 

Ferguson, K. D., and P. M. Erickson. 1987. Will 
it generate AMD? An overview of methods to 
predict acid mine drainage. In Proceedings, 
Acid Mine Drainage Seminar/Workshop, Halifax, 
Nova Scotia, Environment Canada, pp. 215-244. 

Ladwig, K. J., and P. S. A. Campion. 1985. Spoil 
water quality variations at two regraded 
surface mines in Pennsylvania. In 
Proceedings, Symposium on Surface Mining 
Hydrology, Sedimentology, and Reclamation, 
University of Kentucky, Lexington, KY, 
pp. 121-130. 

Lusardi, P. J., and P. M. Erickson. 1985. 

Assessment and reclamation of an abandoned 
acid-producing strip mine in northern Clarion 
County, Pennsylvania. In Proceedings, 
Symposium on Surface Mining Hydrology, 
Sedimentology, and Reclamation, University of 
Kentucky, Lexington, KY, pp. 121-130. 

Smith, R. M., W. E. Grube, Jr., T. Arkle, Jr., and 
A. Sobek. 1974. Mine spoil potentials for 
soil and water quality. U.S. EPA Report 
EPA-670/2-74-070, 303 p. 

Sobek, A. A., W. A. Schuler, J. R. Freeman, and R. 
M. Smith. 1978. Field and laboratory methods 
applicable to overburden and minesoils. U.S. 
EPA Report EPA-600/2-78-054. 

U.S. Environmental Protection Agency. 1983. 
Methods for chemical analysis of water and 
wastes. EPA 600/4-79-020, 490 p. 



19 



RELATIONSHIPS BETWEEN THE INITIAL GEOCHEMISTRY AND 
LEACHATE CHEMISTRY OF WEATHERING OVERBURDEN SAMPLES 



Robert S. Hedin and Patricia M. Erickson 2 



Abstract. — The relationships between geochemical 
measurements and the chemistry of leachates collected in 
weathering tests were explored for 139 overburden samples. All 
samples were analyzed for total sulfur (ST) and neutralization 
potential (NP) and 95 had sulfur fractionated into pyritic, 
sulfate, and organic forms. All samples were weathered for 6 _ 8 
weeks in a laboratory, and leachates were collected and 
analyzed weekly for acidity, alkalinity, and sulfate. All data 
were normalized with logarithmic transformations. 

Cumulative sulfate produced in the weathering tests was 
strongly correlated with total sulfur (r=+.83). The 
relationship with pyritic sulfur was much weaker. NP had a 
negative influence on sulfate production, but the correlation 
was weak (r=-.27), especially for samples with NP less than 30 
tons per 1000 tons. Multiple regression analyses relating 
weekly sulfate production to neutralization potential and sulfur 
forms identified sulfate sulfur as the most important 
independent variable. Pyritic sulfur was not a significant 
factor in any week's analysis. When only samples that had 
produced more sulfate than could be accounted for by sulfate 
sulfur measurements were considered, pyritic sulfur became a 
significant component in the sixth week, and the dominant 
component in the eight week. 

Cumulative acidity/ alkalinity , expressed on a single scale, 
was significantly correlated with total sulfur (r=+.37), NP 
(r=-.H3) and the difference of neutralization and acid potential 
Net NP (r=.6l). Estimates were made of the cumulative 
neutralization generated by each weathered sample by comparing 
sulfate and acidity/ alkalinity production. For most samples, 
less than 10? of the NP had been consumed in the 6-8 week tests. 
All samples with more than 25? consumption produced acidic 
leachates. 



1 Paper presented at the 1988 Mine Drainage and 2 Robert S. Hedin is an Ecologist, Oak Ridge 

Surface Mine Reclamation Conference sponsored by Associated Universities, under contract to U.S. 

the American Society for Surface Mining and Bureau of Mines, Pittsburgh Research Center, 

Reclamation and the U.S. Department of the Interior Pittsburgh, PA, and Patricia M. Erickson is a 

(Bureau of Mines and Office of Surface Mining Supervisory Physical Scientist, U.S. Bureau of 

Reclamation and Enforcement), April 17-22, 1988, Mines, Pittsburgh Research Center, Pittsburgh, PA. 
Pittsburgh, PA. 

21 



INTRODUCTION 

Two procedures commonly used in acid mine 
drainage prediction efforts are acid-base 
accounting and weathering tests. Both procedures 
require that the overburden be adequately sampled. 
With acid-base accounting, the samples are analyzed 
for neutralization potential and sulfur content, 
which is converted into an acid potential. These 
values are compared to determine whether the sample 
is likely to develop acidic or alkaline conditions. 
Samples are often aggregated, by various weighting 
methods, so that the acidity or alkalinity of the 
entire overburden can be represented by a single 
value. 

The acid-base method has several important 
limitations. Sulfur measurements, using current 
procedures, cannot account for the variability that 
exists in sulfide reactivity (Caruccio 1969; 
Hammack 1985; Hammack, elsewhere in these 
procedings). The method used to determine 
neutralization potential can give inflated results 
if the sample contains significant amounts of 
siderite (Caruccio 1967). More importantly, 
acidity and alkalinity are generated by quite 
different processes that are not additive. 
Caruccio and Parizek (1967) have suggested that a 
better approach is to artificially weather samples 
and monitor the chemistry of regularly collected 
leachates. After an appropriate period, samples 
are characterized by their actual production of 
acidity or alkalinity. 

The acid-base method is preferred by most mine 
operators because it has significant cost advantage 
over weathering tests. When overburdens are 
dominated by either acidic or basic materials, 
errors introduced by the acid-base procedure are 
not important. However, when overburdens are 
roughly balanced with respect to acidic and basic 
materials, the errors of the accounting method may 
be unacceptable and weathering tests may be 
required. Despite these additional efforts, post- 
mining water quality is often quite different from 
its predicted composition. 

We evaluated the usefulness of these methods 
for sites with balanced overburden chemistries by 
comparing the results of acid/base accounting and 
weathering tests to measured post-mining water 
quality. In another paper we describe the 
relationships between these prediction methods and 
the drainage chemistry (Erickson and Hedin, 
elsewhere in these proceedings 1988). In this 
paper we discuss the relationship between the 
methods. We try to fit these relationships within 
the current tenets of spoil acidification, and 
where they do not mesh, we identify theoretical or 
methodological problems that cause the failures. 



METHODS 

The overburdens at 32 active m 
described and sampled. At 9 sites 
overburden cores were inspected and 
23 sites fresh highwalls were descr 
samples were collected after removi 
surface materials. Each overburden 
by the thickness and type of indivi 
stratigraphic unit. The entire len 
was sampled and, when this resulted 
unmanageable amount of material, a 
taken. All samples were stored in 
bags. 



ining sites were 
preserved 

sampled. At 
ibed and channel 
ng weathered 

was described 
dual 
gth of each unit 

in an 
subsample was 
sealed plastic 



A total of 264 samples were collected, and a 
determination of total sulfur (ST) (Fisher 
automated sulfur analyzer) and neutralization 
potential (NP) was made for each (Sobek et al . 
1971). (See (table 1) for acronyms of all 
variables discussed). For samples with a total 
sulfur content greater than 0.5? (n=95), sulfur was 
fractionated into organic (SO) sulfate (SS) and 
pyritic (SP) forms using leaching procedures 
described by Sobek et al. (1971). We use the term 
"pyritic" to refer to the sulfur which can be 
extracted by HNO3 but not by HC1. This value is 
believed to represent all forms of reduced sulfur, 
that upon oxidation, would produce acid products. 

Neutralization potentials were expressed in 
tons of CaCOo equivalent per 1000 tons. 3 Sulfur 
contents were expressed as percent of dry weight. 
Acid potentials (AP), expressed in the same units 
as neutralization potential, were estimated by 
multiplying the total sulfur content by 31 .25. Net 
NP was calculated by subtracting acid potential 
from neutralization potential. 

Student's t-test comparisons of the pyritic 
sulfur (t=0.83, P>0.05) and sulfate sulfur (t=1 .30, 
P>0.05) contents of samples collected from 
highwalls and from preserved cores revealed no 
significant differences. Thus, differential 
weathering of highwall and preserved core samples 
was not indicated and all samples were subsequently 
pooled. 

Weathering tests were performed on all 152 
samples from 16 mines. Sample splits of 300 grams 
were crushed to pass a 2mm sieve, and placed in 
plastic boxes which were continuously aerated with 
hamid air for 6-8 weeks. Each sample was initially 
spiked with 10 ml of acid mine drainage to insure 
the presence of iron-oxidizing bacteria. At weekly 
intervals the boxes were filled with 300 ml of 
distilled water and drained after one hour. The 
volume of the collected leachate was determined, 
and the samples were analyzed for pH, sulfate, 
acidity, alkalinity, and iron. 

Weathering test results were expressed as 
milligrams of sulfate or acidity, which were 
calculated by multiplying measured concentrations 
by the volume of leachate collected (table 1). 
Unless otherwise noted, the results of the first 
leachate were not included in calculations. 

Unit sulfate production, (USP), which related 
the cumulative milligrams of sulfate leached to the 
grams of total sulfur present in the sample, was 
calculated: 

USP = SO4 mg/ (sample grams * ST) 

An estimate of the total neutralization generated, 
(NG), was calculated from the stoichiometry of 
pyritic oxidation in the absence of secondary 
neutralization as two moles of H + associated with 
each mole of sulfate: 



NG 



(SO14 mg * 100/96) - acidity 



where acidity was expressed as a positive value, 
and alkalinity was a negative value. 

3The unit derives from agronomic applications, 
where 1000 tons is equivalent to an acre-plow 
layer. Parts per thousand (PPT) is the generic 
term. 



22 



Table 1. — Key to Abbreviations and Calculations. 



Abb. 



Variable 



unit 



Measurement or Calculation 



GEOCHEMICAL MEASUREMENTS 



ST Total Sulfur 
SP Pyritic Sulfur 
SS Sulfate Sulfur 
SO Organic Sulfur 
NP Neutralization 

Potential 
AP Acid Potential 
NETNP Net NP 

Potential 
NP/AP Ratio 



$ of dry wght 
$ of dry wght 
$ of dry wght 
$ of dry wght 

tons/ 1000 tons 
tons/1000 tons 

tons/ 1000 tons 
ratio 



LECO Sulfur Analyzer 1 
acid leaches and LECO 1 
acid leaches and LECO 1 
acid leaches and LECO 1 

titration 1 
ST X 31 .25 

NP - AP 
NP / AP 



LEACHATE MEASUREMENTS 



S04 Cummulative Sulfate 

Production milligrams 

ACID Cummulative Acidity 

Production mg CaCOo equiv 



sum of (sulfate mg/L X L leached) 

sum of ((acidity-alkalinity) mg/L 

X L leached) 
(S04 X 100/96) - ACID 



NG Neutralization mg CaCOo equiv 

Generated 

NC Neutralization $ of NP NG / NP 

Consumed 

USP Unit Sulfate Sulfate per gram S01 / (ST X 300 gm/sample) 

Production of total sulfur 



1 methods from Sobek et al . 1974 

Data were analyzed using the SPSSX computer 
package. We used a 0.05 probability level to judge 
the significance of correlations and regressions. 
Alkalinity was treated as negative acidity so that 
both could be expressed on a single scale. None 
of the data were normally distributed in their raw 
form. Normal distributions of ST, SP, SS, NP, 
USP, NG, and sulfate were obtained with logarithmic 
transformations. Transformations did not normalize 
organic sulfur determinations, so statistical 
evaluations of this minor sulfur fraction should be 
viewed cautiously. Logarithmic transformations of 
acidity values were made after adding 150 to each 
to correct for negative (alkaline) values. For a 
similar reason, Net NP values were increased by 50 
before making logarithmic transformations. 

Because we were most concerned with roughly 
balanced overburdens, we eliminated samples with 
extreme ST or NP values from the analyses. This 
restriction caused the elimination of one sample 
with a total sulfur content of 8.2$ and 12 samples 
with NP > 200. The data set analyzed thus contained 
139 samples for which weathering results, 
neutralization potential and total sulfur were 
known, and a subset of 51 samples for which sulfur 
fractions were also known. 



RESULTS 

The average sample had a total sulfur content 
of 0.64$ and an neutralization potential of 13.3 
tons CaC03 per 1000 tons (n=139). The average 
difference between neutralization potential and 
acid potential (Net NP) was -7.10. Sulfur was 
fractionated for 51 samples which averaged 1.47$ 
total sulfur, of which 77$ was classified as 
pyrite, 17$ as sulfate, and 6$ as organic sulfur. 



The average sample produced in the weathering 
tests a total of 281 mg of sulfate and 98 mg of net 
acidity (calculated: acidity minus alkalinity). 
The 51 samples with fractionated sulfur forms, 
produced on average 624 mg of sulfate and 291 mg of 
net acidity. 



Cumulative Sulfate Production and Acid Potential 

The first hypothesis about relationships 
between leachate chemistry and sample geochemistry 
considered was that sulfate production was related 
to a sample's total sulfur content. This 
expectation was strongly supported (table 2, figure 
1). Regression of acid potential on sulfate 
indicated the following log-log relationship: 

log(sulfate) = 0.675 log(AP) + 1.446 
(r 2 = .77, n=139) 

This estimate of acid potential was based on total 
sulfur values. We expected that acid potential 
values based on pyritic measurements (APSP) would 
result in a stronger relationship. Instead, 
regression analysis revealed a similar equation, 
but a much weaker relationship. 



log(sulfate) = 0.754 log(APSP) + 1 
(r 2 = .35, n-51) 



510, 



The strongest correlation between sulfate 
production and a sulfur fraction was with sulfate 
sulfur (r=.78). Multiple regression of sulfur 
fraction percentages (SP, SO, SS) and NP on 
cumulative sulfate production, produced the 



23 



Table 2. --Pearson correlations between sample 
geochemistry and leachate chemistry. 



Geochemical 
Measurement 



Leachate measurement 
Sulfate Acidity USP NG 



AP 

NP 
Net NP 



+ .37 
- .43 
.55 - . 61 



+ .83 
ns 



AC* 

-.46 

-.27 

ns 



+ .84 
+ .33' 
-.30" 



* P<0.01; *** P<0.001; ns P>0.05 
All variables were log-transformed before 
calculations. 



Caruccio and his students have tested for a 
neutralization effect by comparing unit sulfate 
production (USP) to alkaline production potentials. 
Samples with high NP relative to sulfur, generally 
result in lower production of sulfate per unit of 
sulfur. Caruccio has only discussed a small number 
of samples, all with total sulfur contents > 1$ 
(Caruccio, Geidel and Pelletier 1980). We tested a 
larger number of samples with a wider range of 
sulfur values. Because very low sulfur values 
inflate USP values and the variation in very low 
sulfur values may be instrument or operator 
related, only materials with ST contents greater 
than 0.05? were considered in our analysis. A 
significant, negative correlation was found between 
NP and USP, but it was weak (table 2, figure 2). 



3 
00 



3 

E 

3 

o 




0.01 



0-1 1 10 100 

Acid Potential, Tons CaC03/1000 Tons 



1000 



Figure l.--The relationship between cummulative 
sulfate production by weathering samples and 
their intitial acid potential. Coefficients 
of the regression line are given in the text. 



following relationship. 

log(sulfate) = 0.631 log(SS) + 0.189 log(SO) 
+ 3.^34, (r 2 = .68, n=51) 

Pyritic sulfur and neutralization potential were 
not significant components of the regression 
equation. 



Cumulative Sulfate Production and Neutralization 
Potential 

Several studies have produced results that 
suggest that neutralization potential has a 
negative influence on sulfate production, 
presumably because it inhibits activity by iron 
oxidizing bacteria (Caruccio, et al . 1980; Williams 
et al. 1982, Poissant and Caruccio 1986). We 
tested this hypothesis by comparing NP to 
cumulative sulfate production and unit sulfate 
production. 

NP was not significantly correlated with 
cumulative sulfate production (table 2). This 
finding was further supported by the observation 
that several limestone samples (excluded from these 
analyses) produced very high amounts of sulfate. 
At the sites where these limestones were collected, 
the drainage was characterized by alkaline water 
with extremely high sulfate concentrations. 
Clearly, high NP does not, by itself, cause radical 
reductions in sulfate production. 




1000 



Neutralization Potential, Tons CaCO3/l000 Tons 
Figure 2. --The relationship between unit sulfate 
production and nuetralization potential. 



Cumulative Sulfate Production, AP and NP 

Because both acid potential and neutralization 
potential influence sulfate production, a 
multivariate analysis was used to test the 
complexity of the relationship. Various 
combinations of AP and NP have been suggested as 
appropriate for acid mine drainage prediction 
models. We tested NP, AP, Net NP, and the ratio 
of NP to APST (NP/AP). The best fit was with AP 
and the ratio. 

log(sulfate) = 0.484 log(AP) - 0.215 log(NP/AP) 
+ 1 .608, (r 2 = .80, n=139) 

Note that when NP is less than AP, the ratio is 
less than one and log(NP/AP) is a negative value. 
This results in a higher prediction of sulfate 
production. At ratios greater than 1.0, the 
sulfate production prediction is decreased. This 
relationship indicates that NP only inhibits 
sulfate production when it is present in excess. 

Weekly Sulfate Production by Weathering Samples 

Thus far, only cumulative production of 
sulfate during the weathering tests has been 
discussed. Because leachates were collected and 
analyzed on a weekly basis, analysis of temporal 
changes in the relationships was possible. For 
each week's data, regression equations were 
calculated which related that week's sulfate 



24 



production to the original measurements of 
neutralization potential, pyritic, sulfate and 
organic sulfur. Of the 51 samples analyzed, all 
were weathered for at least 6 weeks, and 45 were 
weathered for 8 weeks. 



Table 4. --Relationships between sample geochemistry 
and weekly sulfate production that are sorted 
according to the proportion of sulfate leached 
to that originally available. 



We predicted that the first week's sulfate 
production would be strongly related to sulfate 
sulfur because during this first leach much of the 
rock's original sulfate content would be dissolved. 
In subsequent weeks we expected that pyritic sulfur 
would be the most important component, because of 
oxidation, and that NP would have a minor negative 
effect on sulfate production. 

Our first prediction was supported by the 
regression analysis (table 3). Sulfate sulfur was 
the most important geochemical component and 
pyritic sulfur was not a significant factor. 
Unexpectedly, sulfate sulfur remained a significant 
part of the regression equations for the entire 
weathering period (8 weeks). Pyritic sulfur never 
became a significant component. Neutralization 
potential acted as a negative factor after four 
weeks. Organic sulfur was a significant factor 
throughout the experiment, but its low degree of 
contribution and its non-normal distribution makes 
specific interpretations difficult. 

Table 3. --Relationships between sample geochemistry 
and weekly sulfate production. The 
coefficients values were calculated for the 
equation: 
S04 = bj(SP) + b 2 (SS) + b 3 (S0) + b 4 (NP) 



a. --Samples with CUMS04/SSAVAIL > 0.80 

S04 = bj(SP) + b 2 (SS) + b 3 (S0) + b 4 (NP) 
Week n r b\ b 2 b 3 b 4 



Week r 


"»1 


b 2 


b 3 


b 4 


1 .69 


ns 


0.70 


0.32 


ns 


2 .77 


ns 


0.72 


0.27 


ns 


3 .84 


ns 


0.71 


0.14 


ns 


4 .82 


ns 


0.66 


0.21 


ns 


5 .86 


ns 


0.64 


0.24 


-0.17 


6 .87 


ns 


0.72 


0.18 


-0.25 


7 .82 


ns 


0.59 


0.24 


-0.24 


8 .75 


ns 


0.56 


0.22 


-0.25 



ns P>0.05; all other coefficients, P<0.05; 
All variables were log-transformed before 
regression analysis. 



To further explore these unexpected findings, 
samples were classified each week according to the 
fraction of sulfate sulfur leached. This fraction 
was calculated by dividing the cumulative sulfate 
leached by the amount of sulfate available in the 
original sample. Samples with a fraction greater 
than 1 .0 had produced more sulfate than could be 
accounted for by the original sulfate sulfur 
content. We tested the hypothesis that sulfate 
production by samples which had been, 
theoretically, depleted of their original sulfate 
sulfur content, was more dependent on pyrite than 
samples for which residual sulfate sulfur still 
existed. 

The highest r-square values for the regression 
equation relating sulfate production to SP, SS, SO 
and NO were obtained when a ratio of 0.80, not 
1 .00, was used as the sorting criteria. We suspect 
that this may be because some of the sulfate sulfur 
present was not readily leached by our weathering 
procedures. The correlations and r-square values 
obtained were consistently higher than those 
obtained in the initial analysis. 



1-3 












4 14 


.84 


ns 


0.76 


ns 


ns 


5 22 


.90 


ns 


0.67 


0.22 


-0.18 


6 26 


.91 


0.52 


0.64 


ns 


-0.36 


7 25 


.89 


0.46 


0.44 


0.15 


-0.33 


8 30 


.84 


0.84 


0.36 


ns 


-0.40 



b. --Samples with CUMS04/SSAVAIL a 0.80 

S04 = bj(SP) + b 2 (SS) + b 3 (S0) + b 4 (NP) 
Week n r b\ b 2 b 3 b 4 



4 


37 


.87 


ns 


0.87 


ns 


ns 


5 


29 


.91 


ns 


0.86 


0.13 


ns 


6 


25 


.94 


ns 


0.98 


ns 


-0.16 


7 


20 


.92 


ns 


0.93 


ns 


ns 


8 


15 


.79 


ns 


0.85 


ns 


ns 



ns P>0.05; all other coefficients, P<0.05 
CUMS04 is the cummulative sulfate produced by 
samples in the appropriate week; SSAVAIL is an 
estimate of the total amount of sulfate contained 
by each sample before weathering; 
All variables were log-transformed before 
regression analysis. 



Sulfate production by samples with less than 
80$ of the original sulfate leached was very 
strongly related to sulfate sulfur and generally 
unrelated to neutralization potential, pyritic or 
organic sulfur (table 4b). The importance of 
sulfate sulfur did not decrease between the fourth 
and eighth weeks for these samples. 

The equations for samples with more than 80$ 
of the original sulfate accounted for in the 
leachate, were considerably different (table 4a). 
Pyritic sulfur was a significant factor after 5 
weeks and the most important factor at eight weeks. 
Sulfate sulfur was a significant component of all 
equations, but its importance decreased with each 
additional week. Neutralization potential was a 
significant negative influence in all but the 
fourth week. 

Two alternative explanations can account for 
the unexpected importance of sulfate sulfur and the 
relative unimportance of pyritic sulfur in 
predicting sulfate production. One explanation is 
that the sulfate sulfur may not have been 
completely solubilized during the first leaching. 
Instead, it could have been gradually released over 
the 6-8 weeks of testing. We find little 
theoretical support for this hypothesis because 
equilibrium and kinetic factors favor complete 
sulfate dissolution. Common sulfate minerals, such 
as gypsum and jarosite, have very rapid dissolution 
kinetics and have high solubilities. All of the 
leachates, which sat for 60 minutes before being 
drained off, were undersaturated with respect to 
these minerals. More importantly, this explanation 
cannot explain the negative influence of 



25 



neutralization potential since it is unlikely that 
calcite or dolomite effect the solubility of 
sulfate minerals or the kinetics of their 
dissolution. It is possible that sulfate is 
contained within particles that are not penetrated 
by the solvent (distilled water) until they undergo 
weathering. We find this process also difficult to 
support. 

Our second explanation is that the sulfate 
sulfur measurement technique is flawed. In our 
method, determinations were made of total sulfur 
for a raw sample and for one which was leached with 
HC1 to remove all sulfate sulfur. Sulfate sulfur 
was calculated as the difference of the two 
measurements. A crucial assumption of this method 
is that HC1 does not effect reduced forms of 
sulfur. Although valid for crystallized, stable 
forms of pyrite, the assumption is not true for 
unstable mono-sulf ides. Soil scientists who work 
with sulf ide-containing sediments commonly use HC1 
to dissolve and quantify reduced sulfur. When we 
perform HC1 leaches on overburden samples, we 
commonly observe hydrogen sulfide fumes, an 
indication of sulfide oxidation (R. Hammack, 
personal communication). This fraction of reduced 
sulfur is the most unstable and most easily 
oxidized. In a weathering experiment, we would 
expect that these sulfur compounds would be the 
first to oxidize and, if present in sufficient 
quantities, significantly effect sulfate 
production during the first several weeks of 
weathering. We would also expect that 
neutralization potential, if present in sufficient 
quantities, could inhibit this source of sulfate. 
Thus, the finding that NP has a significant 
negative role before pyrite has a significant 
positive one is not unexpected. 



*»*J 



-200 



- 1 1 

-150 -100 



-50 



— I — 

too 



Net NP (NP-AP) 



Figure 3. --The relationship between cummulative 
production of net acidity by weathering 
samples and their initial Net NP. 



sulfate levels were high, suggesting that the high 
alkalinity was due to a high amount of 
neutralization. For the other samples with 
unexpectedly high alkalinities, sulfate production 
was low, suggesting limited acid production despite 
high acid producing potentials. 

The best fit regression relating net acidity 
to geochemical measurements was: 

log(Acidity) = -0.364 log(NP/AP) - 0.257 log(AP) 
- 0.802 log (Net NP) + 4.150 
(r 2 = .54, n=139) 



Cumulative Net Acidity and AP and NP 

Net acidity (acidity minus alkalinity) 
produced in leaching tests was significantly 
correlated with both acid potential and 
neutralization potentials, but was most strongly 
related to the net NP (table 2). A plot of net NP 
against net acidity is shown in figure 3. Samples 
with large excesses of neutralization potential 
never produced cumulative alkalinities (negative 
net acidities) greater than 100 mg CaCOo 
equivalent, while several samples with large 
excesses of acid producing potential resulted in 
more than 1000 mg of cumulative net acidity. This 
result is primarily an artifact of the weathering 
experiment and would be expected for all tests with 
a similar design. Constant aeration of the 
weathering boxes prevented the accumulation of 
carbon dioxide and thus limited alkalinities to 
approximately 100 mg over the 6-8 weeks of 
weathering. Even limestones, which are not 
included in Figure 3, never produced more than 200 
mg of alkalinity (40-50 mg/L) . These conditions do 
not represent those commonly observed in spoil 
piles (Lusardi and Erickson, 1985). 



As with sulfate production, a logarithmic model 
best explained production of net acidity, but the 
explained variation was less. This is partially a 
result of the difficulty in fitting a single model 
to data that is influenced by two different 
processes. Generation of alkalinity is limited by 
solubility constraints forced by the gas 
compositions in the weathering vessels. Acidity is 
only limited by the rate of acid potential and, for 
our purposes, is unconstrained by any solubility 
limitations. 

In a second analysis, the influence of 
alkalinity limitations was minimized by only 
considering samples with Net NP less than -8.0. The 
relationship improved considerably and the best 
equation was non- exponential: 



log(Acidity) 



0.004 AP - 0.036 NP 
+ 2.550, (r 2 =.77) 



0.828 NP/AP 



Again, the ratio of neutralization potential to 
acid potential, not the difference, is a 
significant addition to the model. 



All samples, but one, with positive Net NP 
produced alkaline or only slightly acid leachates 
(less than 50 mg CaCO^ equivalent of acidity). 
This relationship held until Net NP reached 
-8.0 tons/1000 tons. Below this value there was 
considerable variation in the relationship between 
Net NP and net leachate acidity. In particular, 
several samples that were extremely deficient in 
neutralization potential (negative Net NP) produced 
alkaline leachates. For two of these samples 



Cumulative Neutralization Generation 

Cumulative generation of neutralization was 
estimated from cumulative sulfate production and 
cumulative acidity/ alkalinity. The calculation is 
based on the assumption that two moles of H + are 
associated with each mole of sulfate. When the 
measured acidity of a leachate sample is less than 
the acidity estimated from sulfate measurements, 
then a calculable amount of neutralization has 



26 



occurred. Generation of neutralization (NG) was 
significantly related with all measures of acid 
potential and neutralization potential (table 1). 
Multiple regression revealed that a logarithmic 
relationship with APST and NP gave the best 
relationship. 



log(NG) = 0.551 log(AP) + 0.138 log(NP) + 
(r2 = .72) 



1 .428 



Because high NG is only possible when a high amount 
of sulfate production has occurred, acid potential 
is more important than neutralization potential. 
Unlike the acidity calculations, separating the 
data into acidic and alkaline overburden samples 
did not improve the relationship. This was because 
NG was less constrained by solubility limitations 
than raw alkalinity measurements. Many alkaline 
samples had high sulfate values, which resulted in 
high NG estimates. 

The above equation also suggests that NP was 
rarely a limiting component of the generation of 
neutralization. To determine the percentage of 
neutralization consumed in the weathering 
experiments, we compared the neutralization 
generated to the neutralization potential 
originally measured. 

Neutralization Consumed = NG mg / NP mg 

Less than 10? of NP had been consumed in 58? of the 
samples. Four of 136 samples had consumption 
percentages greater than 100 indicating that the 
neutralization potential had been theoretically 
exhausted. These samples produced some of the most 
acid leachates. All samples with more than 25J 
consumption produced acidity (figure 4). All of 
these samples had NP values less than 5.0. Samples 
with similar NP values, but lower estimates of 
consumption, generally had alkaline or only 
slightly acid leachates. 




20 40 60 80 100 

Neutralization Consumed, % 



Figure 4. --The relationship between cummulative 
production of net acidity by weathering 
samples and the percentage of NP consumed 
during the weathering tests. 



DISCUSSION 

These results support the general tenets of 
the spoil acidification process. Sulfate 
production by a weathering sample is strongly 
correlated with its sulfur content. The 
relationship is best characterized by logarithmic 
equations. Neutralization potential has an 
inhibitory influence on sulfate production, but the 
effect is variable. It cannot be determined from 
these results whether NP actually inhibits sulfate 
production, or if it is negatively correlated with 
sulfur reactivity. The acidity of a sample is 
strongly related to both its acid producing 
potential (based on total sulfur) and its 
neutralization potential. Both models of sulfate 
and acidity production are improved by the 
incorporation of the ratio of neutralization to 
acid potentials. 

We also found some unexpected results. 
Pyritic sulfur was not a better predictor of a 
leachate's sulfate content or its net acidity than 
total sulfur. This result may be an artifact of 
the limited length of our weathering tests (6-8 
weeks) since, for some samples, pyrite was 
increasing in significance in the final weeks of 
weathering. However, this explanation indicates 
that the results of weathering tests are strongly 
dependent on the period over which they are 
conducted. Because the goal of these tests is to 
obtain an estimate of the acid potential, it is 
important that they are conducted until significant 
pyritic oxidation and acid production has occurred. 
According to Caruccio et al. (1980), the tests 
should be conducted until the rate of sulfate 
production levels off which generally takes 5-10 
weeks. Our results, which are based on obtaining 
a statistically significant relationship between 
sulfate production and the pyritic content of the 
sample, suggest even eight weeks is not sufficient. 
During this study we reviewed an alarming number of 
overburden analyses, done by coal companies or 
consulting firms, which reported leachate results 
after only 3-4 weeks. 

The expected relationship between sulfate 
production and pyritic sulfur only existed for 
samples which were, theoretically, completely 
leached of their original sulfate content and had 
been weathering for at least six weeks. It might 
thus be appropriate to determine the length of a 
weathering test by the ratio of sulfate leached to 
sulfate contained in the original sample. Sulfate 
production rates would not be estimated until more 
sulfate had been leached than had been measured in 
the original sulfate sulfur determinations. 

The unexpectedly strong relationship between 
sulfate production and a sample's sulfate sulfur 
content suggests that sulfate sulfur measurements 
are in error. It is likely that unstable sulfides 
are dissolved in HC1 during the sulfate sulfur 
procedure. Although these compounds may be a small 
fraction of the total pool of reduced sulfur, they 
represent the most reactive forms and would oxidize 
very quickly in weathering tests. If quantifiable, 
this methodological error might be developed into a 
useful indicator of the reactivity of pyritic 
sample. We are currently exploring this 
possibility by analyzing the relationship between 
sulfate sulfur estimates and evolved gas analyses. 



27 



Finally, we address the problem of predicting 
drainage chemistry on a mine level. Are the errors 
calculated and discussed in this paper responsible 
for predictive failures? We think not. The error 
of our relationships is about 25$, which is not 
great enough to account for the large differences 
between expected and actual drainage chemistry 
observed in our study (see Erickson and Hedin in 
these proceedings). Indeed, we appear to have an 
accurate understanding of the acid and sulfate 
production process under controlled, laboratory 
conditions. More likely, the failure of the 
methods lies in conservative assumptions and design 
features that are inappropriate. In the acid- 
base accounting method, the assumption is made that 
all sulfur will oxidize, and all basic materials 
will dissolve. Several surveys of the sulfur 
content of abandoned spoil piles have shown that 
unweathered portions of the spoils retain 
significant quantities of pyritic sulfur (Lusardi 
and Erickson 1985; von Demfange and Warner 1975). 
In our weathering tests, we produced conditions 
that maximized pyritic oxidation and minimized 
alkalinity. Both of these problems can be 
justified when one wishes to make conservative 
judgments. However, if accurate predictions are 
the goal, the assumptions and methodologies must be 
thoroughly reevaluated. 



Sobek, A., W. Schuller, J. Freeman, and R. Smith. 

1971. Field and laboratory methods applicable 
to overburdens and minesolls, 
EPA-600/2-78-05M. 

von Demfange, W., and D. Warner. 1975. Vertical 
distribution of sulfur forms in surface coal mine 
spoils, In proceedings 1975 symposium on surface 
mining, hydrology, sedimentology and reclamation, 
University of Kentucky, p. 135-147. 

Williams, G., A. Rose, R. Parizek, and A. Waters. 
1982. Factors controlling the generation of acid 
mine drainage. Final report on grant G5101086, 
Pennsylvania Mining and Mineral Resources Research 
Institute, Pennsylvania State University, p. 265. 



LITERATURE CITED 

Carucci, F. T., and R. R. Parizek. 1967- An 

evaluation of factors influencing acid mine 
drainage production from various strata of the 
Allegheny Group and the ground water 
interactions in selected areas of western 
Pennsylvania, Coal Research Section, SR-65, 
Pennsylvania State University. 

Caruccio, F. T. 1969. Characterization of strip 
mine drainage, in ecology and reclamation of 
drastically disturbed sites, Gordon and 
Breach, p. 193~224. 

Caruccio, F. T. , G. Geidel, and A. Pelletier. 
1981. Occurrence and prediction of acid 
drainages, in proceedings 1981 symposium on 
surface mining, hydrology, sedimentology and 
reclamation, University of Kentucky, p. 437- 
1*43. 

Hammack, R. W. 1985. The relationship between the 
thermal activity of pyrite and the rate of 
acid generation, in proceedings 1985 symposium 
on surface mining, hydrology, sedimentology 
and reclamation, University of Kentucky, 
p. 139-144. 

Lusardi, P., and P. Erickson. 1985. Assessment 

and reclamation of an abandoned acid-producing 
strip mine in northern Clarion County, 
Pennsylvania, in proceedings 1985 symposium on 
surface mining, hydrology, sedimentology and 
reclamation, University of Kentucky, p. 31 3~ 
323. 

Poissant, S. , and F. Caruccio. 1986. The 

occurrence and viability of thiobacillus 
ferroxidans under varied geochemical 
* conditions in Upshur and Lewis Counties, West 
Virginia, in Proceedings 1986 symposium on 
surface mining, hydrology, sedimentology and 
reclamation, University of Kentucky, p. 141- 
147. 



28 



APPLICATION OF ACID-BASE ANALYSIS TO WASTES FROM BASE 
METAL AND PRECIOUS METAL MINES 1 



Stuart D. Miller and Gavin S. Murray? 



Abstract--Acid-base analysis is used to assess the 
potential for waste materials to generate acid when exposed to 
an oxidized leaching environment. Even though the procedure was 
initially developed for coal mine overburden and reject, it has 
become widely accepted as tool in the assessment of waste rock 
and processing wastes from base metal and precious metal 
operations. 

This paper examines the applications of the acid-base 
analysis to non-coal situations based on the authors' direct 
experience at more than 30 base metal and precious metal 
projects in Australia, New Zealand and Papua New Guinea. The 
fundamental principles of the technique are outlined and the 
major potential pitfalls in interpretation and use of acid-base 
data are described. The paper concludes that the acid-base 
technique is an essential tool for the assessment of waste 
characteristics as well as being suitable as a monitoring 
procedure at non-coal mines. However, the interpretation of 
results is site-specific and requires detailed geochemical 
investigation to provide the understanding and data base for 
this interpretation. 



INTRODUCTION 

The acid-base analysis has been used 
extensively for at least 10 years to identify acid 
forming and potentially acid-forming mining wastes. 
The procedure reported by Sobek et al . (1978) for 
coal mining wastes has been applied in many and 
varied environments, geological materials and in 
many countries. Even though the procedure was 
initially developed for coal mine overburden and 
reject, it has become widely accepted as a tool in 
the assessment of waste rock and processing wastes 
from base metal and precious metal operations. 



!paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of The Interior 
(Bureau of Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

director and Associate, Stuart Miller & 
Associates Pty. Ltd., Sydney, Australia. 



This paper examines the application of the 
acid-base analysis to non-coal situations based on 
the authors' direct experience at more than 30 
base metal and precious metal projects in 
Australia, New Zealand and Papua New Guinea. The 
fundamental principles of the technique are 
outlined and the major potential pitfalls in 
interpretation and use of acid-base data are 
described. 



THE ACID-BASE ACCOUNT: THEORY AND PRACTICES 

The acid-base account, also known as the Net 
Acid Producing Potential (NAPP), is calculated from 
the total sulphur content and the inherent acid 
neutralizing capacity (ANC) of a material. NAPP is 
defined as follows: 

NAPP = %S x 3.125 - ANC 

where ANC and NAPP are in % CaC03 equivalents. 



29 



A negative NAPP indicates that there is excess 
neutralizing capacity and that the material is non- 
acid forming whereas a positive NAPP indicates that 
the material is acid or potentially acid forming. 
The pH and electrical conductivity of a saturated 
paste of crushed (-2mm) material is used to 
support the NAPP classifications. 

Geochemically there are a number of 
fundamental concerns in the theoretical basis on 
which the acid-base or NAPP technique is based. 
The major concerns are as follows: 

1. Assumes all sulphur occurs as pyritic 
sulphide and is acid forming; 

2. Assumes all ANC is available to neutralize 
acid; and 

3. Does not consider the kinetics of either the 
acid forming or acid neutralizing processes. 

The first two concerns may be considered to 
balance each other, however, failure to consider 
the site specific mechanisms and kinetics when 
assessing the likely field geochemistry of a 
particular waste can result in significant error. 

The NAPP procedure is an essential tool in 
waste characterization but the interpretation and 
identification of the implications for waste 
management requires a more detailed determination 
of waste geochemistry. Once the broad principles 
of the site-specific processes have been 
identified, a monitoring and management program can 
normally be developed which is based on the NAPP 
procedure. 

The most important step in carrying out a NAPP 
assessment of waste rock and ore from a base metal 
sulphide or gold operation is sample selection. 
There is generally wide variability in the degree 
of alteration or mineralization associated with a 
mineral deposit. Spatial variability in the 
occurrence of sulphides and carbonates is critical 
and has been responsible for incorrect predictions 
of the acid-forming potential of material when 
composite samples are used in the testing program. 
It is essential that individual depth samples from 
a number of drill holes are tested to provide an 
indication of the variability before composite 
samples are prepared for any detailed geochemical 
analysis. 

APPLICATION OF ACID-BASE ANALYSIS AT SULPHIDE 
MINERAL DEPOSITS 

General 

The objectives of carrying out an acid-base 
analysis are: 

1. Determine the acid-forming 
characteristics of waste rock, ore and 
processing wastes; 

2. Classify waste rock types with respect 
to the potential for acid generation; 

3. Determine the impact on the quality of 
drainage from the mine and waste dumps; 

4. Provide design and management criteria 
for waste disposal, pollution control and 
rehabilitation. 



Our experience at more than 30 base metal and 
precious metal deposits throughout Australia, New 
Zealand and Papua New Guinea has aimed to provide 
the engineers with workable and practical design 
and operational criteria. This includes the 
classification of waste types, disposal needs and 
constraints on the operation. With respect to the 
acid forming characteristics, 4 geochemical types 
are defined as follows. 



CLASSIFICATION 



TRIGGER LEVEL CRITERIA 
pH %S NAPP (%CaC03) 



Acid <4 
Acid Forming <4 
Potentially Acid Forming >4 
Non-Acid Forming >4_ 



<0.5 




>0.5 


>0 




>0 


<0.5 


<0 



In addition to the above 4 types, a 'barren' 
classification is frequently applicable for 
material which is essentially devoid of sulphide 
and acid neutralizing capacity. 

In mineral deposits the mineralogy of 
sulphides and the content of carbonate have a major 
controlling impact on the lag period for acid 
generation and the eventual rate of acid 
generation. Some materials with less the 2% 
sulphur can generate acid in less than 6 weeks 
following exposure, whereas other materials with 
30% sulphur may require 20 years of oxidized 
leaching before acid conditions become established. 

Leaching tests are used to provide design and 
management criteria for controlling acid 
generation. Column and batch leaching tests are 
commonly used, and where possible, larger scale 
field trials are also utilized. Our approach is to 
use coarse (-25mm) size material in columns and 
leach under simulated rainfall conditions. This 
provides an indication of leachate quality, 
weathering behavior, and incorporates the effect of 
the kinetics of the acid formation and acid 
neutralization processes. Batch (humidity cell) 
leaching tests are generally carried out using 
dilute sulphuric acid to neutralize any inherent 
acid neutralizing capacity and create conditions 
under which acid formation can be examined, 
including inoculation with a mixed bacteria 
culture. 

The objectives of the leaching test are to 
provide an indication of the reactivity of 
sulphides, the presence and likely scale of any lag 
period for acid generation, the ability of the 
material to generate acid support bacterially 
catalyzed oxidation, and the nature of short term 
leachates. 

Case Studies 

The variability at any one site can be 
demonstrated by a plot of sulphur and the acid 
neutralizing capacity. Figures 1 and 2 show this 
type of plot for a base metal sulphide deposit in 
Tasmania, Australia (Figure 1) and a gold deposit 
in New Zealand (Figure 2). 



30 



FIGURE I: ACID FORMING POTENTIAL - BASE METAL SULPHIDE DEPOSIT 



at 

3 



(ft 

R 



40 



50 



20 



10 



ACID 
FORMING 



POTENTIALLY ACID FORMING 



^3 — D 



— r - 
10 



~T"" 

20 



-10 

ACID NEUTRALISING CAPACITY XCaCOZ 



NAPP-OKrw 

NON-AC© FORMING 

30 



FIGURE 2 ACID FORMING POTENTIAL - GOLD PROJECT 



10 n 



u 6 



3 
V) 4 

g 



ACID 
FORMWG 



BARREN 






POTENTIALLY ACID FORMING 



NAPP - lirw 



s"~*m no 



out d 
" u ^-" 



NON- ACID FORMING 



5 10 

ACID NEUTRALISING CAPACITY SCaC03 



15 



20 



In the base metal sulphide deposit two 
geochemical waste rock types were identified vis. 
potentially acid forming and non-acid forming. All 
samples contain significant levels of carbonate and 
even the samples with greater than 30% S contain 
between 5 and 20% CaC03 equivalent. Leaching 
studies have been carried out on these rock types 
which confirms the non-acid nature of the non- 
acid forming rock but also confirm the very long 
lag period required for acid generation in the 
potentially acid forming material. It is predicted 
that at least 10 years of oxidized leaching will be 
required before acid is generated in these 
potentially acid forming rocks. Three years of 
operational experience supports this prediction. 
Rock excavated from the underground operation is 
visually assessed as being acid forming or non-acid 
forming with occasional check analysis when new 
areas are opened up. The non-acid material is used 



for engineering works and the acid-forming material 
placed in a controlled stockpile area for later use 
as underground fill in worked out stopes. Mine 
drainage is currently non-acid with a flow rate of 
approximately 30 L/sec. This water is being 
monitored, but it is not expected that acid 
drainage will occur during the life of the mine. 
Final decommissioning will involve sealing and 
flooding of the underground workings. 

Figure 2 shows the acid-base relationship for 
waste rock and tailings from a major gold prospect 
in New Zealand. Four basic waste types are 
identified by this plot as acid forming, 
potentially acid forming, non-acid forming and 
barren. 



31 



The non-acid material has a negative NAPP and 
can be used as general fill or cover for 
engineering works. The barren waste rock is 
essentially devoid of sulphides and also has a low 
ANC. This material plots at the intersection of 
NAPP=0 line and the ANC=0 line. The barren waste 
can be used for any construction purposes providing 
its engineering properties are suitable. 

An important consideration in material 
classified as non-acid is that selective 
dissolution and leaching of carbonates may 
eventually lower the ANC resulting in long term 
acid generation. However, this is only likely to 
occur where the sulphur level is greater than 2 or 
3 percent. 

Integrated geological and geochemical 
investigations show that approximately two-thirds 
of the waste rock and all the tailings fall into 
the acid forming and potentially acid zones. The 
natural pH of the potentially acid forming material 
is greater than 5 but since it has a positive NAPP 
there is a potential for acid generation and 
establishment of low pH conditions (ie. pH less 
than 4) after a period of oxidized leaching. The 
period of oxidized leaching required to generate 
acid conditions increases as the ANC increases (ie. 
as the lag period increases). Also, the potential 
scale of acid generation increases with increasing 
sulphur content. 

Detailed mineralogical and leaching tests 
carried out on the potentially acid forming waste 
material has shown that acid conditions will not 
occur unless the material is exposed to oxidized 
leaching for at least 2 years. At this operation, 
waste materials have been classified according to 
geotechnical and geochemical characteristics. The 
waste disposal strategy has been designed to place 
acid forming and potentially acid forming material 
in cells which are constructed as 'sealed' units 
using non-acid, barren and soft argil ic acid 
forming material as cell walls and cover. The 
soft argil ic material can be used as intermediate 
cover since it can be compacted to high density and 
very low permeability. In addition, the 
mineralogical studies show that the clay matrix 
effectively isolates the sulphides thus minimizing 
exposure to oxidized leaching. 

The tailings generated by processing 
operations in the two gas studies presented are 
classified as potentially acid forming. The base 
metal sulphide tailings contain approximately 30% 
sulphur whereas the gold tailings contain 
approximately 2% S. For control of acid formation 
it is essential that oxidized leaching of the 
tailings is prevented. Both tailings require a 
long period of exposure before acid formation is 
likely to occur and since both sites are in wet 
environments the long term rehabilitation strategy 
is to saturate the tailings and store under 
permanent water cover. The tailings dam walls are 
designed as water-retaining structures and the 
disposal strategies are aimed at providing maximum 
stability to the dam wall. It is considered that 
storage under water is the most effective way of 
preventing long-term acid formation in these waste 
materials. 



CONCLUSIONS 

The approach presented in this paper for 
determining the acid-forming characteristics of 
waste rock, tailings and mine rock has been applied 
to a wide range of base metal sulphide and gold 
mining operations. Acid-base analysis is an 
essential tool in the investigation strategy but 
the interpretation of the results and, most 
importantly, the engineering implications for waste 
management, pollution control and rehabilitation, 
must be carefully considered. Sampling and 
laboratory testing programs must be planned on a 
site-by-site basis and the acid-base analysis must 
be applied on individual samples rather than 
composite samples. 

Once the geochemical waste types have been 
identified at a particular mine operation it is 
essential that a simple monitoring procedure be 
developed for the operation phase. In some 
situations a simple pH determination or visual 
inspection is all that is required. In others it 
may be necessary to determine pH and ANC while the 
complete NAPP and pH determinations may be 
necessary at difficult sites. In any mining 
situation it is generally best to attempt to limit 
the number of waste types to 3 or 4, otherwise 
scheduling difficulties could result in problems 
for reclamation, pollution control and 
decommissioning. 

It is essential to communicate with the design 
and operation personnel since the success of acid 
mine drainage control and prevention of acid soil 
problems for reclamation will depend on the 
incorporation of geochemical factors into the 
scheduling and management of the waste materials. 
Numerous alternatives to the acid-base procedure 
and have been devised and are presented in the 
literature. However, from a practical point of 
view, in relation to the control of acid problems 
through implementation of selective material 
placement and dump design, we believe the acid-base 
procedure along with experienced interpretation is 
an effective and applied approach. 

LITERATURE CITED 

Sobek, A. A., W. A. Schuller, J. R. Freeman and R. 
M. Smith. 1978. Field and laboratory 
methods applicable to overburden and 
minesoils. U.S. EPA Report EPA-600/2-78- 
054. 



A STUDY OF MINE DRAINAGE QUALITY AND PREDICTION USING 
OVERBURDEN ANALYSIS AND PALEOENVIRONMENTAL RECONSTRUCTIONS, 
FAYETTE COUNTY, PENNSYLVANIA 1 



Keith B. 



C. Brady, James R. Shaulis and 
Viktoras W. Skema 2 



Abstract .--The Stony Fork watershed is a 7.44-mi- :: 
drainage basin in Fayette County, PA. Geologic 
investigations during the late 1970 's established that 
surface raining in the watershed was largely limited to 
the Upper Kittanning coal seam. Interpretations of 
drill logs and field work has shown the Upper 
Kittanning and its overburden were deposited in an 
upper delta plain environment. This environment is 
marked by lateral variability and abrupt facies 
changes. This variability affects the distribution of 
calcareous materials, which is evident in acid-base 
account overburden data, and in the water quality 
associated with individual mine sites. Some mines are 
producing alkaline water while others are producing 
acid mine drainage. In addition to the paleo- 
environmental controls on the distribution of 
calcareous strata, it was found that all overburden 
holes less than 40 ft. to the base of the Upper 
Kittanning coal lacked calcareous material. This 
absence is attributed to weathering. Sulfur on the 
other hand persists even at very shallow cover. Mine 
drainage quality is related to the presence or absence 
of calcareous rocks. Surface mining in areas devoid 
of calcareous rocks due to nondeposition or weathering 
has resulted in acid mine drainage. Mines developed 
in areas of abundant limestone produce alkaline water. 



INTRODUCTION 

The Stony Fork watershed, a 7.44-mi 2 
drainage basin, is located in the Fort 
Necessity quadrangle in southern Fayette 
County, Pennsylvania. It is classified as 
a High Quality stream according to 
Pennsylvania Department of Environmental 
Resources (DER) Rules and Regulations and 



L Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and the U.S. 
Department of Interior (Bureau of Mines and 
Office of Surface Mining Reclamation and 
Enforcement), April 17-22, 1988. 



z Keith B. C. Brady is a Hydrogeolo- 
gist, Bureau of Mining and Reclamation, 
Department of Environmental Resources, PA; 
James R. Shaulis and Viktoras W. Skema 
are Geologists, Bureau of Topographic and 
Geologic Survey, Department of Environ- 
mental Resources, Harrisburg, PA. 



has been afforded special protection status 
since the early 1970 's. Because of this 
status, a large amount of hydrologic and 
geologic information has been obtained from 
the applications for surface mines. 
Initially, these data were difficult to 
interpret. Results of overburden and water 
quality testing from mine sites in close 
proximity showed considerable variability. 
These differences were at first attributed 
to the presence of several coals in the 
area having dissimilar overburden 
characteristics. However, attempts at 
correlating these coals became very 
confusing. It soon became apparent that a 
much better understanding of the basic 
geology of the area was needed. 

This was clearly understood by the 
late 1970 's when Pennsylvania's Bureau of 
Forestry and Bureau of Topographic and 
Geological Survey were working in the area 
to develop an estimate of coal resources. 
In order to resolve the coal correlation 
problems, a series of deep exploration core 
holes were drilled. The cores were 
examined in detail, with special care given 



33 



to identifying the key beds in the 
stratigraphic succession, ranging from the 
Conemaugh Group marine beds downward to the 
Mauch Chunk Formation red beds. As a 
result, an accurate stratigraphic framework 
was developed and a better understanding of 
the sedimentological characteristics of the 
area was achieved. It became evident that: 
(1) surface mining in the Stony Fork 
watershed is generally limited to a single 
coal seam, the Upper Kittanning; and (2) 
the variability of mine drainage, both 
alkaline and acid within a relatively short 
distance, is due to complex lateral changes 
in the Upper Kittanning overburden. This 
study attempts to relate the geologic 
features of the area to the guality of 
water produced during mining. It is hoped 
that a better understanding of these 
relationships will enhance predictability 
of potential problem areas and aid in 
preparation of future surface mining plans. 



METHODS 

Forty test holes used for overburden 
analysis were obtained from the Bureau of 
Mining and Reclamation (BMR) surface mine 
permit files. The overburden from these 
holes was analyzed using the acid-base 
accounting method (Sobek and others, 1978) 
which identifies potentially acid and 



alkaline producing strata. Figure 1 shows 
the distribution of the overburden drill 
holes as well as the five most recent areas 
of mining within the watershed. Areas A 
through D have acid-base account overburden 
data. Area E does not have acid-base 
account overburden data, but the geology 
has been studied in detail and it is 
included because of its proximity to the 
other mine sites. 

Values for calcium carbonate equiva- 
lent and percent total sulfur were obtained 
from the overburden analysis test holes. 
The neutralization potential (NP) values 
for each sample were weighted according to 
thickness of each interval. These values 
were then multiplied by the weight of the 
sample stratum per acre foot. The 
resulting value for each hole is in terms 
of tons CaCO^/acre. Likewise the maximum 
potential acidity (MPA) data were summed 
giving tons CaC0 3 /acre deficiency for each 
drill hole. The traditional "excess" and 
"deficiency" (net NP plus MPA) as described 
by Sobek and others, 1978, were not used in 
this study. Included in the calculations 
were all overburden strata above the coal 
and one ft of the strata immediately below 
the coal. Coals were not included. Eight 
of the holes at Site A had NP determined on 
only those strata that "fizzed" when dilute 
hydrochloric acid was applied. Experience 



S\. 



Pennsylvania 



LEGE 



ROAD 

STREAM 

U.K. CROPLINE 

40 FT. OVERBURD! 
THICKNESS LINE 

MINING LIMIT 
DRILL HOLE 
WATER SAMPLE P" 




Figure l.--Map of study area showing overburden analysis drill hole locations and associated mine sites 
(A through E). The numbers beside the drill hole locations indicate the amount of calcareous material 
in the hole presented as thousands of tons of CaC03 per acre. Note that all holes outside the 40 foot 
cover line have 0.0 amount of calcareous material. 



34 



has shown that rocks with NP's less than 30 
do not normally exhibit a fizz, whereas 
rocks with NP's greater than 30 normally 
do. In order to include all data in our 
calculations and to make the data 
internally consistent, only NP's greater 
than 30 with a fizz were used. Experience 
has shown that NP's less than 30 do not 
generally contribute significant alkalinity 
and total sulfur values of less than 0.5 
percent often do not result in poor guality 
water. Only sulfur values greater than 0.5 
percent and neutralization potentials 
greater than 30 tons CaCO 3 /1000 tons were 
used in the calculations. A comparison 
between these data versus data including 
all NP and sulfur values yielded similar 
results . 

Water sample data used in this study 
were obtained from BMR permit files. The 
analyses include both BMR samples and 
company monitoring data. 

Geologic cross sections and isopach 
maps were constructed to determine the 
thickness distribution and lateral and 
vertical relationships of the various rock 
units in the area. Additionally, a series 
of maps was drawn showing the areal 
distribution of rock types at different 
stratigraphic levels. These slices through 
the strata at arbitrary vertical positions 
are intended to show successive patterns of 
sediment deposition occurring throughout 
the area at different instances of time. 
This method of recreating contemporaneous 
depositional environments is an approx- 
imation. Differences in compaction of 
dissimilar sediments, downcutting of stream 
channels and original topographic relief 
are factors which distort precise time line 
reconstructions . 






EXPLANATION 


. — . 


DRftlNAGl IASIN BOUMDAftV 


— 


MINI till I0UND*RT 


• 


o*U rOIHT 


.— , 


LINES OF CROSS SICTI0K 




SCALE 




o moon 







Figure 2. — Data and section line location map. 
Bold letters (A-E) designate mine sites 
discussed in text. 



relationship is construction of a strati- 
graphic framework. This was accomplished 
by correlating the persistent marine zones 
found in the lower part of the overlying 
Conemaugh Group. The Brush Creek, Pine 
Creek, and Woods Run marine zones are 
persistent throughout the area and were 
recognized in every drill hole that 
penetrated their horizons (fig. 5). The 
first red beds encountered in the upper 
part of the Mauch Chunk Formation also 
provided useful stratigraphic control from 
below. 



A wide range of types of data was used 
for developing the geologic cross sections 
and maps. Reliability of information used 
was generally very good. Most of it 
consisted of geologists' descriptions of 
cored exploration drill holes, overburden 
test hole logs, and surface mine highwalls. 
This was supplemented by the somewhat less 
reliable descriptions of rock cuttings from 
air rotary drill holes submitted with min- 
ing permit applications. The data points 
used in the construction of cross-sections, 
isopach maps and lithology distribution 
maps are shown in figure 2. 

GENERAL GEOLOGY 

The stratigraphic section examined in 
this study consists of the stratigraphic 
interval from 20 ft below to 100 ft above 
the Upper Kittanning coal. It ranges from 
the calcareous rocks associated with the 
Johnstown limestone at the bottom of the 
section, to the rocks overlying the Upper 
Freeport coal horizon at the top (figs. 3 & 
4). The prominent feature throughout this 
section is the laterally discontinuous 
nature of the rock units. Correlation of 
these units is difficult, and an essential 
first step in understanding their lateral 



Sedimentologically the prominent 
features in this study are the Upper 
Kittanning coal, a sandstone overlying the 
coal, and several calcareous beds. The 
Upper Kittanning Coal is generally thick 
and continuous over most of the area (fig. 
6). Several splits of the coal are 
apparent in section Y-Y' (fig. 3). One 
occurs in the area of holes 7 and 11. The 
coal here is split by a sandy siltstone 
wedge which becomes thicker and coarser to 
the west. On the southern end of section 
Y-Y' (Hole 8) a 2.5-ft argillaceous 
limestone parting in the coal suggests 
splitting in a southward direction. 

The sandstone overlying the coal in 
the western part of the drainage basin has 
a lenticular, elongate shape. The thicker 
part of this lense is approximately a mile 
wide in the south and narrows to less than 
a half mile wide northward (fig. 7). The 
sandstone is coarse-grained at the base and 
becomes fine-grained upward. The thick, 
main body of the sandstone grades laterally 
into thin, very fine-grained sandstones 
interbedded with sandy silt shales, silty 
clay shales, and calcareous silty 
claystones. 



35 




EXPLANATION 
L T L \ Calcareous Rock 



10000 Feet 



[*• / 1 Sandstone 

Q£] Siltstone 
I ICIaystone 



Figure 3.--Stratigraphic cross section (Y-V) of the Upper Kittanning to Upper Freeport interval. 

* z' 

A-2 A-l 




I I I I | I I I I | 
O lOOOO F es t 



Figure 4.--Stratigraphic cross section (Z-Z 1 ) of the Upper Kittanning to Lower Freeport interval. 



This part of the section also has 
three, discrete, major calcareous zones. 
These zones often contain argillaceous 
limestones and contain freshwater fossils 
( Spirorbis and ostracodes). The most 
persistent and thickest of these is beneath 
the Upper Freeport coal horizon 
approximately 50 to 80 ft above the Upper 
Kittanning coal (figs. 3 & 4). Another 
thick but discontinuous zone is present 
below the Upper Kittanning in the western 
part of the area. This calcareous zone 
attains a maximum thickness of 25 ft and 
usually contains the argillaceous Johnstown 
limestone. The area where the limestone is 
thickest and least argillaceous coincides 



with the thickest development of the Upper 
Kittanning coal and overlying sandstone. 
South of the Stony Fork drainage basin this 
calcareous zone appears to be laterally 
eguivalent to the Upper Kittanning coal. 
Limestone in drill hole 8 is found directly 
below and above the coal and forms a thick 
parting in the coal itself (fig. 3). A few 
small isolated pods of limestone directly 
on top of the coal were encountered at 
other locations in and around the basin. A 
third major calcareous zone is situated 
approximately 15 to 40 ft above the Upper 
Kittanning coal. Its position varies from 
just below the Lower Freeport coal horizon 
to midway between the Lower Freeport and 



36 



U*«r Biktrstm 




Figure 5.--Stratigraphic cross section of the Mauch Chunk Formation to Lower Bakerstown coal interval 
See figure 2 for location of section. 



Upper Kittanning. 

In addition to the well-developed 
Upper Kittanning coal, two other coals of 
minor significance are present in and 
around the drainage basin. The Lower 
Freeport coal has a sporadic distribution 
in the study area. It reaches a maximum 



thickness of 2 to 3 ft, but is thin to 
absent over much of the area. The Upper 
Freeport coal is rarely present in the 
study area. Carbonaceous clay shales, 
calcareous claystones, and argillaceous 
limestones containing fossil ostracodes and 
Spirorbis often occupy its horizon. 



37 





Figure 6. --Thickness map of the Upper Kittanning 
coal. Thickness lines given in feet. 



Figure 7. --Thickness of the sandstone above 
the Upper Kittanning coal. Thickness 
lines given in feet. 



PALEOGEOGRAPHY 

The map showing distribution of 
lithologies 1 ft below the Upper Kittanning 
coal indicates the presence of a large body 
of sandstone in the eastern portion of the 
study area (fig. 8). Not enough infor- 
mation is present to determine its gross 
geometry, however a few deep holes indicate 
that the sandstone which is as much as 40 
ft thick grades upward from coarse to very 
fine-grained. It is adjacent to a narrow 
belt of claystone to the west which 
separates the sandstone from a large area 
of calcareous claystone and argillaceous 
limestone containing fossil freshwater 
invertebrates. These calcareous deposits 
are indicative of a freshwater lake or pond 
environment. The narrow belt of claystone 
and adjacent very fine grained upper part 
of the sandstone body suggest the presence 
of overbank deposits flanking a levee and 
stream channel system to the east. 

This period of time appears to be 
followed by a nearly complete cessation of 
deposition over the entire study area, and 
the subsequent development of luxuriant 
vegetation of the Upper Kittanning peat 
swamp. Some local ponding of water too 
deep to sustain vegetation occurred in the 
swamp complex. This resulted in the 
formation of the limestone partings seen in 
several places. The coal split located in 
the area of drill holes 7 and 11 may 
indicate the presence of a brief very local 
incursion of sediment into the peat swamp 
from the north-west (fig. 3). Closer 
spacing of drill holes is needed in this 
area to better understand this minor 
episode of deposition. 



Distribution of lithologies 10 ft and 
25 ft above the coal provide a good look at 
the nature of the sediment influx which 
terminated Upper Kittanning peat devel- 
opment (figs. 9 and 10). Initially a 
blanket of sand and silt was deposited over 
the peat throughout most of the area. A 
fluvial channel oriented roughly north- 
south (fig. 7) and associated levee de- 
posits were soon established in the area 
underlain by the thickest coal and 
limestone. In the early stages of sediment 
influx over the peat swamp this area of 
thick peat would probably have undergone 
the most compaction and would have quickly 
localized the channelway. As a result, 



EXPLANATION 
| I | i| Calcareous Rock 

[311 Sandstone 

|"H Siltstone 

| | Claystone 




Figure 8. --Distribution of lithologies one 
foot below the Upper Kittanning Coal. 
Letters designate mine site locations. 



38 



EXPLANATION 

H-rH Calcareous Rock 

Sandstone 
1 | Siltstone 

| [ Claystone 




Figure 9.— Distribution of lithologies 10 feet 
above the Upper Kittanning coal. Letters 
designate mine site locations. 




EXP LANATION 
P^-L| Calcareous Rock 

|".- .• | Sandstone 

| ~~| Siltstone 

| Claystone 




FT NECESSITY 7 Vj 



5000 



FEET 



Figure 10. --Distribution of lithologies 25 
feet above the Upper Kittanning coal. 
Letters designate mine site locations. 



ponded overbank areas of very low energy 
were established soon after flooding of the 
swamp. Calcareous claystones, such as the 
ones found in the overburden of mine sites 
A and C were being deposited in close 
proximity to the high energy channel system 
(fig. 9). The area between mine sites A 
and C just to the east of the channel 
developed into a persistent site of low 
energy (figs. 10 and 11). By the time the 
sediments 50 ft above the coal were 
deposited the main channel area appears to 
have lost much energy and become very 
narrow (fig. 11). The major episode of 
ponding which was to cover the entire area 
with calcareous deposits by Upper Freeport 
time, was just beginning in the western 
part of the study area. 

The map of rock distribution 10 ft 
above the coal shows an area in the 
northeastern part of the Stony Fork 
watershed in which thin coal is present 
(fig. 9). This is an area where 
considerable splitting of the Upper 
Kittanning coal has been 
mining. It is just west 
containing relatively thick 
grained sandstone with some granular size 
grains. This could indicate the presence 
of a channel system contemporaneous with 
the peat swamp to the east. This channel 
system may be related to the sandstone body 
found directly under the coal in the 
eastern portion of the study area (fig. 8). 

In summary, considering the relatively 
narrow width of the stream channel system 
and adjoining levees, and the preponderance 
of freshwater lake and pond sediments 
throughout the section, this area was most 
probably part of an interdistributary area 
of an upper delta plain during Upper 
Kittanning to Upper Freeport time. Except 



observed in 
of an area 
very coarse 



for some clays, these areas received little 
sediment input; as a result aggradation 
could not keep up with subsidence, and 
small water bodies were formed. The 
development and enlargement of these ponds 
drowned out the peat-producing marshes and 
swamps. Streams carrying a sandy bed load 
in this type of environment can be expected 
to be small and sinuous (Frazier and 
Osanik, 1969). As a result patterns of 
deposition within this environment are 
complex with many lateral changes. 



EXPLANATION 
Calcareous Rock 

Sandstone 

Siltstone 




Figure 11. --Distribution of lithologies 50 
feet above the Upper Kittanning coal. 
Letters designate mine site locations. 



39 



DISTRIBUTION OF NEUTRALIZATION POTENTIALS 
AND SULFUR AS THEY RELATE TO WEATHERING 

In this area there is a weathering 
profile to a depth of 20 feet below the 
surface which has selectively affected the 
calcareous rocks and pyrite concentrations 
of the Upper Kittanning overburden in the 
study area. Field observations and 
drilling records have reported brown to 
orange coloration of the rocks and soil in 
this interval. The effects of weathering 



on CaCO: 



can be seen in 



the NP data 
gathered'for the study area. A plot of the 
NP data in terms of tons CaC03 per acre ft 
versus thickness of overburden on the Upper 
Kittanning coal seam in figure 12 clearly 
shows a CaC03 threshold at a thickness of 
about 40 ft. Holes that encountered less 
than 40 ft of Upper Kittanning overburden 
lack calcareous material. Holes that 
penetrated more than 40 ft of Upper 
Kittanning overburden show a wide range of 
values, from to hundreds of tons 
CaC03/acre ft. There are two reasons for 
this. In the first place, test holes in 
the study area show that significant 
occurrences of limestone are usually found 
20 ft or more above the Upper Kittanning 
coal seam (figs. 3 & 4). In the second 
place, weathering has leached all the 
carbonate material to a depth of 20 ft 
below the surface. Because of these two 
factors, more than 40 ft of Upper 
Kittanning overburden is required before 
any measurable amounts of CaC03 can be 
found. If the effects of weathering were 
eliminated overburden thicknesses of only 
20 ft could have calcareous material 
present. Figure 13 is a plot of sulfur 
versus thickness of Upper Kittanning 
overburden ( sulfur is expressed in terms of 
tons CaC03 deficiency/acre ft). Unlike 
the calcareous material, the sulfur data 
shows no overburden thickness threshold. 
Sulfur greater than 1.0 percent was present 
at depths of only 10 to 15 ft below the 
surface. It appears that weathering of 
CaC03 is more complete than weathering of 
pyrite at shallow depths. Therefore, areas 
within the Stony Fork watershed with less 
than 40 ft of overburden will lack 
calcareous material, but still potentially 
have high sulfur concentrations. 

Figure 12 shows that when there is 
more than 40 ft of overburden there is a 
wide range in available alkaline material. 
We attribute this to the regional and 
patchy distribution of freshwater limestone 
and occasional fluvial channels cutting out 
and replacing the limestone with non- 
calcareous rocks. 



300- 








i • 


r- 










o 










o 








• 


L. 










UJ 








• 


or 










O200- 








' * 


< 








! * • 










• 


X 








: 


o 








• 


o 





















o 








• 


100-i 








!• * * 


V) • 








z 










o 








• 


H 








i 










! . • • . _ 


0- 


IIIIIIMIIIIIlfl 


[Hermit 


I iftlffltl 1 1 |l t 1 'II . .' | ■ . r ] 1 [ i i i 1 1 | I 1 1 ■. . ■ 1 1 1 | | 1 1 


6 io 


20 


30 


40 SO 60 70 80 90 100 



DEPTH TO UPPER KITTANNING COAL 

Figure 12. --Graph showing relationship between 
calcareous material per acre foot and 
depth to bottom of Upper Kittanning coal, 
calculations were made based on 
neutralization potentials above 30 tons/1000 
tons. 



O 

o 

U. 
LJ 

or 
o 
< 

> 
o 

z 
LJ 

o 
u_ 

LJ 

a 

IO 

o 
o 
o 
o 



80 n 



60 



20- 



*• • 



• • • * • 



n M 1 1 1 1 n i ) 1 1 1 1 1 | f u i ) 1 1 1 1 1 1 1 1 1 1 1 1 n 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 n n 1 1 li n i n 1 1 1 1 H n 1 1 1 1| l n 1 1 1 1 1 ) | n 1 1 1 1 1 1 M 

10 20 30 40 50 60 70 80 90 100 

DEPTH TO UPPER KITTANNING COAL 

Figure 13. --Graph showing relationship between 
potentially acid producing strata and 
depth to bottom of Upper Kittanning coal. 
Calculations were made based on sulfur 
values above 0.5%. 



thickness. Calcareous rocks where present 
consisted of either argillaceous limestone 
or claystone with limestone nodules. The 
maximum highwall heights ranged from a low 
at site D of 45 ft to a high at sites A and 
C of 90 ft. 



WATER QUALITY OF MINE SITES 

In each of the five mine sites studied 
(A through E) the Upper Kittanning coal was 
the principal seam of interest. The Lower 
Freeport coal horizon was also encountered 
at each site but generally appeared as a 
thin coal or a carbonaceous shale except at 
site C where it reached a minable 



Determinations of water quality were 
made on pit water samples taken during 
mining, and on water from crop springs, 
monitoring wells, and spoil seeps taken 
both during and after mining. These 
samples were analyzed for pH, alkalinity, 
acidity, iron, manganese, aluminum, and 
sulfate. Summaries of selected parameters 
are shown in tables 1 and 2 . 



40 



TABLE 1. -PIT WATER DURING ACTIVE MINING 



Mine 
Site 


N 


pH 


Nel Acidity! 
mg/1 


Sulfate 
mg/1 


Med. 


Q12 


93 


Med. 


ei 


93 


Med. 


91 


93 


A 


14 


67 


6.2 


7.0 


-68 


•85 


•34 


101 


43 


166 


B 


6 


6.0 


5.4 


66 


16 


■15 


59 


17 


6 


29 


C 


2 


7.2 


7.1 


7.3 


•154 


•224 


•84 


79 


59 


99 


D 


19 


3.0 


2.0 


3.1 


406 


250 


1028 


630 


420 


900 


E 


9 


3.3 


3.1 


3.8 


158 


65 


194 


380 


250 


592 



• Net acidity = acidity - alkalinity. Negative number indicates net alkalinity. 
2 Q1 and Q3, upper and lower quartilies determined by sorting 







TABLE 2. - 


POSTMINING WATER QUALITY 






Sample 
Point 


N 


pH 


Net Acidity' 
mg/1 


Sulfate 
mg/1 


Med. 


Q12 


93 


Med. 


91 


93 


Med.3 


91 


93 


SPA 










see f 


Sure 14 


graph 








SP-B 


4 


5.6 


5.5 


5.8 


6.0 


0.2 


11.8 


<43 






SP-C 


7 


6.0 


5.7 


6.3 


-4 


•34 


4 


<40 






SP-D 


9 


3.2 


3.0 


4.2 


456 


293 


609 


918 


756 


1040 


SP-E 


4 


2.8 


2.7 


2.9 


988 


796 


1154 


1281 


1112 


1545 






INet acidity = acidity - alkalinity. Negative number indicates net alkalinity. 

2 Q1 and Q3, upper and lower quartilies determined by sorting. 

3 Less than symbol, " < " indicate that some lab values" reported as " <40 mg/1" S0 4 



Mine site A was activated in early 
1985 and was nearing completion at the time 
of this writing. It occurs in an area of 
paleoenvironmental transition. The mine 
site encompasses both the coarser clastic 
sediments associated with the distributary 
channel system located in the west and the 
lower-energy overbank sediments, including 
limestones to the east. The effects of 
this transition are reflected in the 
overburden analyses and in the water 
quality associated with this site. Three 
alkaline zones were encountered at 2-10 ft, 
20-25 ft, and 50-65 ft above the Upper 
Kittanning coal. This operation had a 
maximum overburden height of 85-90 ft. 
These zones occur as lenticular bodies that 
are very limited in their lateral extent. 
The lowest alkaline zone was an 8 ft thick 
lenticular parting at the top of the Upper 
Kittanning coal occupying an area of only 
about 20,000 ft 2 . Some test holes starting 
70 ft above the Upper Kittanning coal had 
little or no alkaline material present. 
Analysis of pit water samples have been 
consistently alkaline, although sulfate 
showed a general increase through time 
indicating the presence of both alkaline 
and acid-producing rocks. A down gradient 
monitoring well below the earliest 
backfilled area has begun to show a rise in 
sulfate and metals (fig. 14). The precise 
cause of this will require further 
investigation, however, some possible 
explanations include an insufficient amount 
of alkaline material present to inhibit 
pyrite oxidation or more acid-producing 
material present than anticipated. Further 
monitoring will be necessary to determine 
whether the increases in concentration are 
a persisting trend or a short term 
fluctuation. 

Mining at site B was initiated in July 
1983 and was completed by October 1984. 
The geology at mine site B is similar to 



mine site A. The depositional environment 
of the overburden rocks in the western most 
part of the site was the same distributary 
channel system found in site A. Overburden 
in this area consisted primarily of 
sandstone for a distance of 50 ft above the 
Upper Kittanning coal. Eastward the 
sandstone thinned abruptly and silts, clays 
and calcareous sediments were deposited in 
a low energy overbank environment. Two 
main alkaline bearing zones were found 
15-20 ft and 50-60 ft above the Upper 
Kittanning Coal seam which was mined to a 
maximum highwall height of 70 ft. The two 
alkaline zones were found to be persistent 
throughout the site except in the extreme 
western edge. The pit water contained low 
concentrations of sulfate and metals and 
ranged from weakly acidic to weakly 
alkaline. Sample SP-B (table 2) from a 
crop spring is weakly acidic with low 
concentrations of sulfate and metals. The 
pit water at mine site B and the crop 
spring sample SP-B show virtually no impact 
from mining. Normally postmining water 
quality will show an increase in dissolved 
solids. The water quality in the deeper 
portions of the mine site where the 
rock/water contact time is greatest may not 
be reflected at sample point SP-B. It may 
be that SP-B only reflects shallow flow 
from the periphery of the mine site. 

Mine site C was first started in 
January 1985 and is active at the time of 
this writing. Mine Site C is located 
almost entirely in a low energy deposi- 
tional environment marked by extensive 
deposition of limestone and calcareous 
strata. As in site A three alkaline 
bearing zones were present at 5-10 ft, 
20-25 ft and 50-60 ft above the Upper 
Kittanning coal seam, and the site also had 
a maximum overburden of 85 to 90 ft. 
Unlike site A however, the upper two 
limestone zones were found to be laterally 
persistent over the entire site. Also, 
unlike site A or B, the Lower Freeport coal 
was of sufficient thickness to be a minable 
seam. The pit water was highly alkaline 
with low sulfate and metals concentrations 
due to the abundance of alkaline rock 
strata present. A low flow seep, SP-C 
(table 2), developed below a backfilled 
portion of the site, is characterized as 
mildly alkaline, low sulfate, with metals 
concentration slightly above background. 
Because this site is still active, a full 
evaluation of postmining water quality is 
not possible and will not be until the 
ground water regime reestablishes itself. 
Seep SP-C probably represents relatively 
shallow flow of short retention time. Of 
the sites discussed this mine will 
encounter the largest quantity of 
alkaline-bearing strata due to the 
persistence of the calcareous zones and the 
large area above 40 ft of cover. It is 
anticipated that the alkaline material will 
be sufficient to prevent postmining water 
quality problems. 

Mine site D was activated in late 
1979, with mining on the decline by late 



41 



2000 




10/64 11/84 4/86 5/86 6/86 8/86 7/87 10/87 



Figure 14. --Graph of Mine Site A monitoring 
well sulfate concentrations through time. 



quality with low pH and high acidity and 
sulfate (table 1). Toe of spoil seeps 
display similar but somewhat poorer water 
quality characteristics (sample SP-E, table 
2 ) . In addition a pre-existing Upper 
Kittanning country bank deep mine 500 ft 
south of SP-E also showed a deterioration 
in water quality during active mining 
(Stump and Mastrilli, 1985) and continues 
to be degraded. 

SUMMARY AND CONCLUSIONS 

Based on the foregoing study of the 
Stony Fork watershed, the following conclu- 
sions can be made: 

1) The distribution of alkaline 
material is the primary control 
on mine drainage quality. Areas 
lacking alkaline strata produce 
acid mine drainage; areas with 
significant alkaline strata pro- 
duce alkaline water. 



1981. The site has now been abandoned and 
has been left unreclaimed. Mine site D is 
located in a high energy depositional 
environment. This is suggested by the 
presence of thick coarse grained sandstone 
overlying the coal, a lack of calcareous 
strata and numerous claystone partings and 
boney layers in the coal. No alkaline 
bearing rocks were found in the overburden 
which reached a maximum height above the 
Upper Kittanning coal of 40 to 45 ft. In 
this site the Upper Kittanning coal 
uncharacteristically contains numerous 
carbonaceous, pyritic claystone partings 
sometimes as great as 2 ft in thickness. 
The Lower Freeport coal was generally thin 
to absent over the entire site except for a 
several acre area where it reached a 
minable thickness. Pit waters during 
mining were highly acidic with high sulfate 
and metals concentrations (table 1). Seeps 
which developed along the toe of spoil are 
of even poorer water quality. This quality 
of water is attributable to the lack of 
calcareous material and the indiscriminate 
spoiling of substantial quantities of 
carbonaceous, pyritic claystone partings. 

Mine site E was actively mined from 
1979 to 1981, when it was abandoned; it 
remains unreclaimed. The overburden rocks 
of mine site E were deposited in the 
north-south trending distributary channel 
system seen also on the western edges of 
mine sites A and B. Thick massive sand- 
stones deposited in the main channel and 
steeply dipping interbedded siltstones and 
sandstones associated with point bars are 
the predominant rock types found over much 
of the site. These high energy environment 
deposits persist throughout the overburden 
as high as 8 5 ft above the Upper Kittanning 
coal. No alkaline-bearing zones were found 
in the Upper Kittanning overburden. As in 
Site D the Lower Freeport coal was 
generally thin or absent, except in a very 
small area where it became thick enough to 
be minable. Pit water was always of poor 



2) Areas where the Upper Kittanning 
overburden is less than 40 ft 
thick are devoid of alkaline ma- 
terial due to weathering of the 
alkaline zone that occurs approx- 
imately 20 feet above the coal. 
High sulfur content, however, 
occurs within strata as shallow 
as 10 to 15 ft from the surface. 
Therefore, a zone from the crop- 
line to a 40 ft overburden 
thickness line lacks alkaline 
strata, but sulfur bearing strata 
still persist within this zone. 
Consequently, mines developed in 
overburden of 40 ft or less will 
most likely produce acid mine 
drainage. 

3 ) The Upper Kittanning and Lower 
Freeport coals and overburden 
were deposited in an upper delta 
plain interdistributary environ- 
ment. This environment is marked 
by narrow stream channel systems 
and freshwater lakes and ponds. 
Mine sites developed in or ad- 
jacent to the high energy paleo- 
stream channels lack calcareous 
materials. Coals in these areas 
often contain claystone partings 
and boney layers. Mine sites 
located in low energy areas away 
from the channels are marked by 
extensive calcareous deposits. 
Because of this variability an 
upper delta plain environment 
does not necessarily ensure good 
quality water. 

4) Mines developed in areas of high 
energy depositional environments 
are acid mine drainage producing 
whereas mines developed in areas 
of low energy depositional envir- 
onments are alkaline producing. 
The exception to this would be 
mines of low cover in low energy 
depositional environments that 



42 



would be acid producing due to 
weathering of calcium carbonate. 

5) In areas of significant paleo- 
environmental variability, adja- 
cent raining cannot be relied upon 
as a mine drainage quality 
predictor. Site-specific over- 
burden analyses are important for 
delineating the distribution of 
alkaline and acidic materials. 

6) Pit water quality in a broad 
sense appears to be a fairly good 
indicator of postmining water 
quality, although pit water quali- 
ty is generally lower in metals 
and sulfate concentrations. 

7) Mining plans should be designed 
such that sufficient alkaline 
material will be encountered. In 
the Stony Fork study area 
thickest Upper Kittanning coal 
development tends to coincide 
with areas which have the 
thickest sandstone overburden and 
which lack alkaline material. 
However, the coal in these areas 
tends to be underlain by thick 
limestone. A possible mining 
plan in these cases would involve 
deepening the pit in order to 
incorporate the underlying lime- 
stone into the spoil. In other 
instances when alkaline material 
does occur in the overburden, but 
only under high cover, the cuts 
could be run perpendicular to 
contour in order to blend the 
alkaline material evenly through- 
out the backfill. 



ogical Survey, 4th ser. 
Resources Report 91, 88 p. 



Mineral 



Sobek, A. A., W.A. Schuller, J.R. Freeman, 
and R.M. Smith, 1978, Field and 
Laboratory Methods Applicable to 
Overburdens and Minesoils. U.S. EPA 
EPA-600/2-78-054, 203 p. 

Stump, D.E. and T.M. Mastrilli, 1985, 
Effects of Surface Mining on Streamflow, 
Suspended-Sediment, and Water Quality in 
the Stony Fork Drainage Basin, Fayette 
County, Pennsylvania. U.S. Geological 
Survey Water Resources Investigations 
Report 84-4362, 28 p. 



ACKNOWLEDGMENTS 

Susan M. King assisted with computer 
calculations and construction of a 
structure contour map, Richard G. Dudginski 
assisted with computer graphics. Michael 
W. Smith provided useful comments and 
assistance throughout the project. The 
manuscript benefited from comments and 
review by Roger Hornberger and Tim Kanai. 
Teresa Smith typed the manuscript. 



LITERATURE CITED 



Frazier, D.E. and A. Osanik, 1969, Recent 
peat deposits - Louisiana coastal plain, 
in Dapples, E.C., and Hopkins, M.E. 
(eds.), Environments of Coal Deposition, 
Geological Society of America Special 
Paper 114, pp. 63-85. 

Shaulis, J.R., 1985, Coal Resources of 
Fayette County, Pennsylvania; Part 1. 
Coal Croplines, Mined-out areas, and 
Structure Contours. Pennsylvania Geol- 



43 



PHOSPHATIC CLAY SLURRIES FOR REDUCING ACID MINE DRAINAGE 
FROM RECLAIMED MINE SITES 

Eric D. Chiado and Dr. John J. Bowders and Dr. John C. Sencindiver * 



Abstract. — Research at West Virginia University is being 
conducted on developing a phosphatic clay slurry seal that will 
prevent or reduce acid mine drainage produced from reclaimed 
sites. The spoil material that is used to reclaim sites in West 
Virginia consists primarily of shale and sandstone. When 
exposed to oxygen and water the spoil produces acid mine 
drainage. By developing an effective phosphatic clay slurry 
seal, it will be possible to hydrologically isolate the acid 
producing materials. An added benefit of this application is 
that phosphatic clay, which is a waste product of the 
production of fertilizer, will be utilized, thereby reducing the 
need for storage of these clay wastes. An effective way of 
hydrologically isolating the acid producting material is to 
reclaim sites using a layered system consisting of spoil 
material, overlain with a phosphatic clay slurry which is in 
turn overlain by cover soil. Current research is being 
conducted to define an optimum system that will produce the 
lowest permeability and reduce the acidity of any effluent 
generated in the acid materials. To date, laboratory 

permeability test results indicate that the compactive effort is 
the dominant variable in determining the permeability of the 
spoil-slurry system. Low water content slurries (150% to 250%) 
should be applied to uncompacted spoils, and high water 
content (250% to 350%) slurries should be applied to compacted 
spoils. The thickness of the slurry does not appear to 
significantly alter the permeability of the spoil-slurry system. 
The addition of the slurry has resulted in a 40 to 80 percent 
decrease in the permeability of the spoil material. The addition 
of slurry also dramatically decreases the concentrations of iron, 
manganeese, magnesium, aluminum, and sulfates in the effluent. 
Slight increases in the pH of the effluent were also 
demonstrated. Additional testing is underway to establish the 
optimal conditions for obtaining minimum permeabilities. 



^Eric D. Chiado is a Graduate Research Assistant in Civil 
Engineering, Dr. John J. Bowders, Jr. is an Assistant Professor of 
Civil Engineering and Dr. John C. Sencindiver is an Associate 
Professor of Plant and Soil Science, at West Virginia University, 
Morgantown, WV. 



44 



In the past 50 years an ever increasing 
amount of research has been conducted on the 
control and remediation of acid mine drainage 
(AMD) produced from both deep and surface coal 
mines. Many of these effortB have concentrated on 
the use of crushed limestone or other base 
materials to neutralize the AMD. These methods, 
however, required perpetual reapplication to insure 
sufficient buffering capacity. Puthermore, the 

application of base materials could only neutralize 
the acid, but was unable to stop or prevent the 
acid producing reaction. The increasing 

environmental awareness of the public and the 
increasing cost of remediation has made it desirable 
for the coal mining industry to seek alternative 
methods for controlling AMD. 

To this end, research is being conducted at 
West Virginia University on the use of phosphatic 
clay to control the formation of AMD. Phosphatic 
clay is the waste product of the production of 
fertilizer. Approximately 400 million tons of this 
waste are currently stockpiled in Florida, and this 
amount is expected to exceed 1 billion tons by the 
year 2000 (Lin). Using this waste product for the 
control of AMD will have the added benefit of 
reducing some of the disposal problems facing the 
phosphate fertilizer industry. 

Phosphatic clay is made up of approximately 
25% apatite, 60% smectitic clay minerals, and 15% 
various other phosphate minerals (Renton). The 
apatite consists of calcium phosphate (Ca5(PC>4)3), 
which becomes readily soluble below a pH of 7.0. 
This dissolution frees the phosphate ion, which has 
a strong affinity for metal ions. The smectitic clay 
minerals are composed primarily of montmorillonite. 
When exposed to water, montomorillonite expands 
several times its original volume, which results in a 
very low permeability. These two characteristics of 
phosphatic clay; the ability to release phosphate 
ions and low permeability, make it extremely 
attractive for use in controlling the formation of 
AMD. 

It is generally accepted that the oxidation- 
hydrolysis of pyrite (FeS2) contained in spoil 
material and coal refuse produces AMD. When 
pyrite is exposed to oxygen and water, it 

dissociates into ferrous ions (Fe + 2) and sulfate ions 
(SC>4 - 2). The ferrous ion is futher oxidized to the 
ferric ion (Fe + 3). Once the ferric ion has been 
formed it takes over as the primary oxidizing 
agent, and these reactions do not stop until the 
pyrite supply is exhausted (Renton). As these 
reactions progress metals such as magnesium and 
aluminum, if present, can also be leached from the 
spoil. The end result from all of this chemical 
activity is AMD. The quality of AMD can range 
from low acidity with high concentrations of 
sulfates and low concentrations of metals, to AMD 
high in acidity, metals, and sulfates (Kleinmann). 

The application of phosphatic clay to reclaimed 
mine sites has the ability to not only neutralize, 
but also completely stop this reaction (Renton). An 
efficient way to achieve this goal is to place a 
slurry of phosphatic clay over the potentially acid 
producing spoil. The montmorillonite provides a 
first line of defense in the form of a hydraulic seal 
against infiltrating surface water. Because of 



montmorillonite's low permeability, any surface 
water that is not eventually evaporated will require 
a great deal of time to pass through the seal. Any 
water that does manage to pass through the seal 
will react with any exposed pyrite. The resulting 
acidic environment will dissolve the calcium 
phosphate. The phosphate ions will react with the 
ferric and ferrous ions to form almost insoluble 
iron phosphates. The major oxidizing agent will be 
removed, and the AMD producing reaction will be 
stopped (Renton). 

The objective of this research is to determine 
those variables most important in producing 
phosphatic clay slurry seals that possess low 
permeability. The information gained from these 
experiments will be used to develop field scale 
tests, which will in turn, be used to make 
recommendations for coal operators to apply 
phosphatic clay seals to toxic materials, thus 
reducing AMD. 

MATERIALS AND METHODS 

Samples of phosphatic clay were obtained 
from a phosphate mine located near Mulberry, 
Florida. The phosphatic clay was shipped in 
watertight 55-gallon containers. Samples of spoil 
material were obtained from an active surface mine 
located in central West Virginia. Both of the 
materials are indicative of the general physical and 
chemical nature of materials located in the vicinity 
from which they were obtained. A cover soil was 
also obtained from the same reclaimed area as the 
spoil. This cover soil will be applied over the 
phosphatic clay seal to prevent the subsequent 
drying and cracking of the seal that will occur if 
the seal is exposed to the atmosphere. The cover 
soil will also provide a vertical overburden stress 
or weight on the underlying seal and spoil, and it 
will also provide a base upon which to grow 
vegetation. 

Several standard tests were performed on 
each seperate material to determine its physical 
characteristics. All tests were conducted in 

accordance with ASTM standards (ASTM, 1985) and 
consisted of : hygroscopic water content 
(D2216-80), specific gravity (D 854-83), standard 
proctor compaction <D 698-78), Atterberg limits 

(D4318-84), grain size analysis (D 422-63), and 
hydrometer analysis (D422-63). These tests allowed 
extrapolation of the results to similar materials not 
included in this research. Results from these tests 
are shown in Table 1. 

The next step in the research involved 
developing mixing procedures for the preparation 
of slurries. Slurries ranging in water content from 
150% to 375% by weight were prepared from the 
phosphatic clay. A prerequisite for acceptance of 
a slurry at a certain water content for futher 
testing was that it had to have the ability to be 
poured so that it could be applied in the field. 

Before the clay was mixed with water, it was 
passed through a funnel with a half-inch diameter 
opening. Clay clods larger than this opening were 
broken by hand. The clay was then hand mixed 
with a spatula to negate any gradation that may 
have taken place. This procedure was followed to 



45 



insure reproducible results, and to provide clay 
clods that could be placed in suspension by 
standard laboratory equipment. Initially a 

slow-action two-speed mixer was used to prepare 
the slurries. The clay however, tended to clump 
into large clods, and was very plastic. A 

rapid-action mixer, run at high speed was then 
used, and this produced smooth slurries. This 
mixer was used throughout the remainder of the 
research to prepare the slurries. 

The major portion of the research consisted of 
permeating various configurations of spoil and 
slurry. A 10.5 cm diameter by 11.5 cm high rigid- 
wall double-ring permeameter (Figure 1) was used 
for all of the tests. The bottom of this 

permeameter consists of two concentric porous 
stones separated by a steel ring. The diameter of 
the ring is such that the area of the inner and 
outer porous stones are equal, so that under ideal 
conditions, the outflow is divided equally between 
the two porous stones. The inner ring is isolated 
from the sides of the permeameter, and therefore, 
side wall leakage, which produces inflated values 
for permeability, does not affect this inner stone. 
All values for permeability were calculated using 
the inflow to the permeameter. Outflows from the 
inner and outer stones were also recorded to make 
sure no side wall leakage occurred. Side wall 
leakage was indicated whenever the outflow from 
the outer stone was much greater than the outflow 
from the inner stone. 

Distilled deionized water was used as the 
permeant liquid, and this was stored in a 
nonreactive acrylic reservoir. Pressure was 

supplied to the reservoir by a pressure board, 
which forced the water out of the reservoir into 
the permeameter and thereby set up a hydraulic 
gradient through the vertical axis of the specimen. 
The hydraulic, gradients used in the test ranged 
from 10 to 200 (cm/cm). Gradients for each test 
were determined solely on their ability to expedite 
test results. 

Permeability tests were run separately on the 
spoil and slurry to provide a base upon which to 
later evaluate the effect that various combinations 
of spoil and slurry had on permeability. At least 
two permeability tests were run on specimens of 
each material to make sure that results were 
reproducible. Spoil material specimens were 

permeated in both the uncompacted and compacted 
states, and slurries ranging in water content from 
150% to 375% by weight were also tested. 

With the completion of these "base" 
permeability tests, experiments were conducted to 
test the permeability of layered spoil-slurry 

arrangements. These tests were done to determine 
the effects of slurry water content, slurry 
thickness, and degree of spoil compaction on the 
permeability of the layered systems. Tests were 
conducted using: 1?) loose spoil (no compactive 
effort), 2) compacted spoil (100% standard Proctor), 
3) high water content slurries (>275%), 4) low water 
content slurries (<275%), 5) thick slurries (>3cm), 
and 6) thin slurries (<3cm). Eight permeability 
tests were required to cover all possible 
arrangements of these variables. See Table 2 for 
the testing setup. Each layering arrangement was 



permeated for at least four pore volumes of flow. 
One pore volume is equal to the volume of voids in 
the sample. This insured complete saturation of 
the voids in the sample, and provided a 
nondimensionalized means of comparing results from 
different specimens. Upon completion of the test 
each specimen was extruded from the permeameter. 
Settlement of the specimen was measured, and any 
distinquishing features such as cracking of the 
slurry and rust deposition from the permeater were 
recorded. The specimen was then split along its 
vertical axis in several places, and the penetration 
of the slurry into the spoil was measured. 

During the permeability tests, effluent from 
several specimens was collected for chemical 
analysis. This analysis included determinations of 
pH, alkalinity-acidity, and concentrations of iron, 
manganeese, aluminum, calcium, magnesium, sulfates, 
and phosphates. Basic elemental constituents were 
analyzed using atomic absorption, pH was 
determined using a digital ionalyzer, sulfates were 
analyzed using the turbidometric method, and 
colorimetric analysis was used to determine 
phosphate concentrations. These analyses were 
done to determine the quality of effluent produced 
by the spoil, and to determine what effect the 
addition of phosphatic clay slurry had on effluent 
quality. 

RESULTS AND CONCLUSIONS 

Typical behaviors of uncompacted and 
compacted spoil specimens both before and after 
the addition of slurry are shown in Figure 2. All 
of the uncompacted specimens demonstrated 
decreasing permeability before the addition of 
slurry because they were found to have settled. 
This impeded the flow of water, which resulted in 
the decreasing permeabilities. The compacted 

specimens exhibited steady to slightly increasing 
permeability before the addition of slurry because 
these specimens had swelled. This enabled the 
water to move through larger flow paths, which 
resulted in the increasing permeabilities. Before 
the addition of slurry every specimen exhibited 
erratic shifts in permeability between individual 
permeability readings. After the slurry was added 
an immediate decrease in permeability occurred, 
and the variation in permeability between 
individual readings stabilized somewhat. After the 
initial large decrease, the permeability of the 
uncompacted specimens tended to decrease, whereas 
the permeability of the compacted specimens 
remained the same or increased slightly. 

Results from the "base" permeability tests and 
the eight spoil-slurry permeability tests are shown 
in Table 2. Before the addition of slurry, the 
average permeability of the uncompacted spoil was 
approximately 1.8 x 10~4 cm/s, and that of the 
compacted spoil was approximately 1.2 x 10"^ cm/s. 
Tests 2(1) through 2(4), performed exclusively on 
slurry, show it to possess an aveage permeability 
of approximately 3.0 x 10"^ cm/s. 

Based on this last result it was expected that 
spoil-slurry arrangements would achieve 

permeabilities similar to 3.0 x 10~6 cm/s. Tests 
3(1) through 6(2) show that this is not true. 
Permeabilities for these tests ranged from a high 



46 



Table 1. Index Properties for Spoil Material and Phosphatic Clay 



Soil 


Natural 

Water 
Content 


Plastic 
Limit 


Liquid 
Limit 


Specific 
Gravity 


Maximum 

Dry Unit 

Weight 

(lb/fO 


Optimum 

Water 
Content 


Spoil 


12.5 


24 


31 


2.27 


105 


13 


Clay 


31.5 


35 


135 


2.80 


75 


12 



Table 2. Soil arrangement and permeability results for "base" and 
spoil-slurry permeability tests. 



Test 




Slurry 

Water 

Content 


Slurry 
Thickness 


Spoil 
Avg. K 


Spoil 

and 

Slurry 

Avg. K 


Change 
in K 


No. 


Arrangement 


(*) 


(cm) 


(cm/s) 


(cm/s) 


w 


1(1) 


Loose Spoil 


NA 


NA 


1.67xl6 4 


NA 


NA 


1(2) 


Loose Spoil 


NA 


NA 


1.4 xlO 4 


NA 


NA 


Z(l) 


Slurry 


179(L) 


5.56(Th) 


*3.63xl0" 7 


NA 


NA 


Z(2) 


Slurry 


194(L) 


7.18(Th) 


*1.83xl0~ 6 


NA 


NA 


2(3) 


Slurry 


329(H) 


2.61(T) 


*4.61xl0~ 6 


NA 


NA 


2(4) 


Slurry 


389(H) 


4.6(Th) 


*5.24xl0~ 6 


NA 


NA 


3(1) 


Compacted 
Spoil & Slurry 


274 (L) 


3.10(Th) 


2.09xl0~ 5 


6.12xl0~ 6 


70.7 


3(2) 


Compacted 
Spoil & Slurry 


353(H) 


3.71(Th) 


2.20xl0~ 5 


5.95xl0~ 6 


73.0 


4(1) 


Loose Spoil & 
Slurry 


174(L) 


3.43(Th) 


2.69xl0~ 4 


2.66xl0" 5 


90.1 


4(2) 


Loose Spoil & 
Slurry 


225(L) 


3.93(Th) 


1.53xl0~ 4 


6.93xl0 _5 


54.7 


5(1) 


Compacted 
Spoil & Slurry 


187(L) 


1.77(T) 


3.57xl0~ 6 


3.93xl0" 6 


-10.1 


5(2) 


Compacted 
Spoil & Slurry 


300(H) 


2.54(T) 


4.08xl0~ 6 


3.30xl0~ 6 


19.1 


6(1) 


Loose Spoil 
& Slurry 


187(L) 


1.27(T) 


2.69xl0~ 4 


9.50xl0 _6 


96.5 


6(2) 


Loose Spoil 
& Slurry 


300(H) 


2.1KT) 


1.53xl0~ 4 


2.77xl0~ 5 


81.9 


(H) 


= high water cc 


>ntent 










(L) 


= low water coi 


itent 










(Th) 


= thick slurry 












(T) 


= thin slurry 












K 


= permeability 













*For tests 2(1) through 2(4) 
the slurry. 



the permeability listed in column five is for 



47 



PRESSURE 



00 1000 



0.000100 

PERMEABILITY 
(CM/3fC) 

0,000010 



0.000001 



2 VENT 

gjp__a 

7-r\\t i i\\i / i. i r r\ 




INNER 
OUTLET 



OUTER 
OUTLET 



Figure 1 - Double Ring Permeameter 



1 . - 1 


* Loose Spoil/Slurry 

w-1878 Slurry Applied * 
3.95 PV 

Compacted Spoil/Slurry 
w-27455 Slurry applied Q 
6.FPV 


- 




* * . 


£ 


* & * 


** 


1 








i 




A o o 




■- 




-■'- 




• ■ - 













10 



Figure 2- Typical Results From the Permeability Tests on Spoil Material and 

Phosphatlc Slurry 



48 



of 7 x 10~5 cm/s to a low of 3 x 10~6 cm/s, with an 
average of 2 x 10 ~5 cm/s. All of these tests have 
average permeabilities higher than that of the 
slurry by itself. It appears from this result that 
the spoil plays an important part in determining 
the overall permeability of spoil-slurry systems. 
The reason for this result is not readily 
discernable. Variables used in tests 2(1) through 
2(4), such as hydraulic gradient and slurry water 
content, were similar to the range of values used 
in tests 3(1) through 6(2). The thicknesses of the 
slurries used in tests 2(1) through 2(4) were 
approximately 2 times thicker than slurries used in 
tests 3(1) through 6(2). This difference however, 
still does not explain why the permeability of the 
slurry does not govern the overall permeability 
because test 2(3), with a thickness of 2.61 cm, has 
a permeability lower than six out of the eight 
spoil-slurry tests, which had similar slurry 
thicknesses. 

The results in Table 2, with the exception of 
test 5(1), show that the addition of slurry will 
decrease the permeability of the spoil. The 

magnitude of this decrease, shown in the column 
labelled percent change in permeability, is not 
constant. On the average, a 38% decrease in 
permeability, with the addition of slurry, can be 
expected for compacted spoil, and an 81% decrease 
in permeability, with the addition of slurry, can be 
expected for loose spoil. 

Table 3 lists the eight spoil-slurry specimens 
and their respective final permeability, percent 
change in permeablility with the addition of slurry, 
and overall performance. Each test was ranked 

from one to eight with respect to the final 
permeability, with one representing the lowest 
permeability, and eight representing the highest. 
The tests were also ranked from one to eight with 
respect to the change in permeability, with one 
representing the largest change, and eight 
representing the smallest. The percent change in 
permeability was calculated by dividing the 
difference between the average permeability of the 
spoil alone and the average permeability for the 
spoil-slurry system by the average permeability of 
the spoil alone 

Table 3 demonstrates that the specimens with 
the lowest final permeability also tended to exhibit 
the smallest change in permeability when the 
slurry was added. Because compactive efforts in 
the field are often suspect, and uncompacted or 
undercompacted spoil may be the existing condition, 
it was equally important to develop spoil-slurry 
systems that produced low final permeabilities, 
and large changes in permeability with the addition 
of slurry. Table 3 allows for the determination of 
the best possible methods for applying spoil-slurry 
systems. By adding the point values from the final 
permeability column and the change in permeability 
column it is possible to rank the overall 
performance of different combinations of spoil and 
slurry. Lower valves of overall performance 

represent better systems. 

From Table 3 it can also be seen that the 
lowest permeabilities were achieved when the spoil 
was compacted. The reason for this is that 

compaction reduces void spaces, and flow paths 



become smaller and more tortuous. Tests on 

uncompacted spoil had higher permeabilities 
because flow paths were not as constricted, and 
were straighter. As was stated previously, the 
permeability of the overall system is dependent on 
both the spoil and the slurry, and not the slurry 
alone. The addition of slurry can only do so much 
to lower the permeability of the system. The 
remaining decrease in permeability comes from the 
compactive effort applied to the spoil, with 
increasing effort producing lower permeability. 

For the compacted specimens the water 
content of the slurry did not seem to have as 
much of an effect on the permeability as did the 
thickness of the slurry. This result may be 
misleading however, because the large range in 
permeabilities between the compacted specimens 
before the addition of slurry makes it dificult to 
judge the effect of slurry water content and 
thickness on the permeability of the compacted 
spoil. The overall performance ranking of the 
compacted specimens provided a less biased means 
of evaluating the effect of these variables on the 
permeability. Using this ranking, it appears that, 
in fact, the water content of the slurry was more 
important than the slurry thickness in determining 
the final permeability of the compacted spoil-slurry 
specimens. Higher water content slurries, whether 
thick or thin, produced better overall performances 
in the compacted specimens than did lower water 
content slurries. The reason for this may be that 
the higher water content slurries can penetrate the 
compacted spoil and totally or partially clog the 
flow paths, thereby futher hindering water 
movement. The lower water content slurries may 
not have been fluid enough to penetrate the 
compacted surface. Compacted specimens split 

after permeability testing showed no visible bulk 
penetration for either low or high water content 
slurries. The clogging of these flow spaces 

occurred on the microscopic level however, so 
blocked flow paths may not have been discernable 
to the naked eye. 

The range in permeability between the 
uncompacted specimens before the addition of 
slurry was small, and therefore, correlation 
between final permeability, change in permeability, 
and overall performance was much better. As was 
the case with compacted specimens, the water 
content of the slurry appeared to have a greater 
effect on the permeability than did thickness of 
the slurry. In this case though, lower water 
content slurries produced better overall 

performances in the uncompacted spoil specimens. 
Because the uncompacted spoil had large flow 
paths, the higher water content slurries may have 
been flushed out of the specimens instead of 
clogging the flow paths. Lower water content 
slurries were less fluid, and may have provided a 
more stable barrier against the flow of water. 

A final conclusion that can be drawn from 
Table 3 is that the best overall performance was 
produced from an uncompacted spoil specimen 
overlain with a low water content slurry. Attempts 
at compaction should still be made because 
compacted specimens demonstrated the lowest 
permeabilities, however, it is still possible to 
achieve statiBfactory results using uncompacted 



49 



spoil as long as the slurry covering the spoil has 
a low water content. 

Results from the chemical analysis of effluent 
are shown in Table 4. Analyses before and after 
the addition of slurry were only performed on test 
specimens 3(1) and 3(2). The pH of the effluent 
for both specimens increased slightly with the 
addition of slurry, however, it did not reach 
neutrality. The reason for this result is that 
phosphatic clay is a relatively neutral material (pH 
= 7.2) so that it is difficult for it to completely 
buffer an acidic material. 

The affinity of phosphate ions for metal ions 
is dramatically demonstrated in the large 
reductions in the concentrations of iron, 
manganeese, magnesium, and aluminum, with the 
addition of slurry. The absence of any detectable 
concentration of phosphate ions after the addition 
of slurry shows that the phosphate ions were 
completely consumed in various chemical reactions 
with the metal ions to produce insoluble metal 
phosphates. A visual inspection of the color of the 
effluent also indicated that the phosphate ions 
were effective in removing the metals. The yellow 
color of the effluent, common in AMD, was apparent 
before the addition of slurry, however, with the 
addition of slurry the effluent became clear within 
one pore volume of flow. 

The increase in the concentration of calcium 
with the addition of slurry is due to the 
dissolution of the calcium phosphate contained in 
the phosphatic clay. This result leads to an 
increase in the hardness of the effluent, however, 
this consequence must be weighed against the 
advantages of the large removal of heavy metals. 



content slurries ranging from 250% to 350% by 
weight should be applied to compacted spoils. The 
thickness of the slurry does not seem to be 
important in determining the permeability of the 
spoil-slurry system. It may be advantageous, 

however, to use thick slurries to better insure that 
any cracks that may develop in the seal do not 
pass completely through the seal. Before the 
addition of slurry, uncompacted spoil material has 
an average permeability of 1.8 x lO - '* cm/s. After 
the slurry has been applied the average 
permeability can be expected to drop to 3.3 x 10~5 
cm/s. Before the addition of slurry, compacted 
spoil material has an average permeability of 1.3 x 
10 - 5 cm/s. After the slurry has been applied the 
average permeability can be expected to drop to 
4.8 x 10~6 cm/s. Spoil material should be 

compacted before the application of slurry. 
Relatively low permeabilities can still be produced 
in uncompacted spoil if the slurry water content is 
kept low. The addition of slurry alBO greatly 
reduces the concentrations of iron, manganeese, 
magnesium, aluminum, and sulfates in the effluent 
through the formation of insoluble metal 
phosphates. The concentration of calcium in the 
effluent will increase due to the dissolution of 
calcium phosphate in the slurry. All of the various 
chemical reactions will result in a slight increase in 
the pH of the effluent with the addition slurry. 

Results from these tests will be applied to 
field tests during the summer and fall of 1988. It 
is hoped that these test will futher enable the 
authors to define the optimum spoil-slurry system, 
and allow them to develop procedures for applying 
this technology to the coal mining industry. 



Sulfate concentrations were also reduced, but 
the final concentrations were still high. The 

reductions in sulfate concentration of the two 
specimens are reflected in the increase of pH for 
both specimens. 

A discrepency in the iron concentration 
existed for test specimen 3(2). The initial 

concentration was very low, but increased 
significantly with the addition of slurry. An 

explanation for the initially low concentration may 
be that the sample of spoil used to prepare the 
specimen did not contain a representative amount 
of pyrite. The increase in iron concentration after 
the addition of slurry may be due to outside 
sources of iron that were introduced to the 
specimen during testing. An inspection of the 
permeameter after the permeability test was 
completed indicated that spalling of the interior 
wall had occurred. This could have provided a 
large amount of iron that flowed down the interface 
between the permeameter and the specimen. 

In conclusion, it appears that the compactive 
effort applied to the spoil is the dominant variable 
in determining the permeability of spoil-slurry 
systems. Larger compactive efforts result in lower 
permeabilities. Slurry water contents are also 
important in determining the permeability of 
spoil-slurry systems. Low water content slurries 
ranging from 150% to 250% by weight should be 
applied to uncompacted spoils, and high water 



ACKNOWLEDGEMENTS 

The investigators gratefully acknowledge the 
support of this project by Project Officer Dr. 
Charles R. Jenkins of the Water Research Institute 
(Project No. 08). The assistance of Daniel L. Rogers 
and D. Todd Mooney with the laboratory tasks is 
also acknowledged. 



LITERATURE CITED 

1. Lin, K. LT., Figueroa, J. I., and Chang, W. F., 
"Engineering Properties 

of Phosphogypsum." Proceedings of the 
Second Workshop on By-Products of 
Phosphate Industries. Miami, Florida, 

April 1984. 

2. Renton, John, J., "Acid Mine Drainage", 
Mountain State Geology, 

1985, Boyd Press, Wheeling, WV, pp 1-6. 

3. Kleinmann, R.L.P., and Erickson, P.M. "Control 

of Acid Mine Drainage: An Overview of 
Recent Developments" Proceedings of the 
National Mined Lands Reclamation 

Conference, St. Louis, MO, Oct 28-29, 

1986, pp. 283-305. 

4. ASTM, Annual Book of Standards, Vol. 4.08, Soil 
and Rock; Building Stones Philadelphia, PA, 1985. 



50 



Table 3. Ranking of spoil-slurry tests for the lowest final permeability, greatest change 
in permeability, and best overall performance, based on results from Table 1 



Test No. 



Soil Arrangement 



Final 
Permeability(l) 



Change in 
Permeability(2) 



Sum of 
Columns 
3 and 4 



Overall 
Performance 



(4) 



3(1) 
3(2) 

4(1) 

4(2) 

5(1) 

5(2) 

6(1) 
6(2) 



Compacted Spoil 
& Slurry 

Compacted Spoil 
& Slurry 

Loose Spoil & 
Slurry 

Loose Spoil fe 
Slurry 

Compacted Spoil 
& Slurry 

Compacted Spoil 
& Slurry 

Loose Spoil & 
Slurry 

Loose Spoil & 
Slurry 



14 



10 



10 



(1) 1 = lowest final permeability 8 - highest final permeability 

(2) 1 = largest change 8 = smallest change 

(3) Numbers determined by adding the final permeability ranking to the 
change in permeability ranking 

(4) Overall performance determined by ranking numbers in column five from 
lowest to highest, with lowest being the best. 



Table 4. Results of Effluent Analysis Before and After the Addition of 

Phosphate Clay Slurry 



Test 




P H 


Fe 


Mn 


Ca 


Mg 


Al 


so" 2 


-4 3 


3(1) 


Before Slurry 


3.89 


81.9 


248 


249 


557 


623 


9320 


0.25 


After Slurry 


5.73 


5.6 


17 


658 


80 


6 


2145 


— 


3(2) 


Before Slurry 


4.14 


13 


271 


300 


586 


345 


8580 


0.25 


After Slurry 


4.64 


160 


107 


603 


78 


13 


2040 


— 



Concentrations are in parts per million 



51 



LABORATORY AND FIELD TESTING OF A 
SALT-SUPPLEMENTED CLAY CAP 
AS AN IMPERMEABLE SEAL OVER PYRITIC SLATES 1 



Eric W. Murray, S. Paul Goudey, Ron G. L. McCready and Joe Salley^ 



Abstract — During construction of a taxiway at 
Halifax International Airport in Nova Scotia in 1982, 
about 225,000 cubic metres of fractured waste pyritic 
slate material was disposed in a pile covering about 
7 hectares on the airport property. As a result of 
air and water infiltration and the presence of 
Thiobacillus f errooxidans , this waste rock pile has 
been seeping a heavy metal laden acidic effluent 
similar to acid mine drainage. Because sufficient 
quantities of a relatively impermeable clay 
overburden were present on site, it was decided to 
apply a compacted clay cap to the pile to reduce air 
and water infiltration. From previous experience it 
was known that concentrations of greater than 1% 
sodium chloride in the aqueous phase are inhibitory 
to Thiobacillus f errooxidans . Laboratory studies 
confirmed that addition of a layer of highway salt 
beneath the clay not only inhibits the production of 
acid drainage, but also enhances the sealing capacity 
of the clay. Placement of the salt supplemented clay 
cap was undertaken in 1986 and 1987. Early 
observations indicate a reduction in the volume and 
an improvement in the quality of water seeping from 
the waste rock pile. 



INTRODUCTION 

Halifax International Airport in Nova 
Scotia, Canada is located on a band of 



1 Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation Con- 
ference sponsored by the American Society 
for Surface Mining and Reclamation and the 
U.S. Department of the Interior (Bureau of 
Mines and Office of Surface Mining Reclama- 
tion and Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

o . . 

' Eric W. Murray is a Project Manager 

with Public Works Canada, Moncton, N. B. , 
S. Paul Goudey is Environmental Officer 
with Transport Canada at Halifax Interna- 
tional Airport, Ron G. L. McCready is Head 
of the Biotechnology Section of the Canada 
Centre for Mineral and Energy Technology 
(CANMET) in Ottawa, Ontario, and Joe Salley 
is a Physical Chemist at CANMET in Ottawa. 



highly mineralized pyritic slate bedrock. 
As part of construction of an aircraft 
taxiway in 1982, a mass of this slate was 
blasted and excavated. About 225,000 cubic 
metres of waste slate material was deposit- 
ed in a pile covering about 7 hectares ad- 
jacent to the taxiway. Several months af- 
ter the start of construction an effluent 
of acid water laden with heavy metals began 
seeping from the waste rock pile and from 
the subdrains beneath the taxiway. At the 
direction of regulatory authorities, a sys- 
tem was immediately constructed to collect 
this acid drainage and to treat it by addi- 
tion of hydrated lime. At the time, it was 
thought that the acid drainage would sub- 
side in several months. 

When the acid seepage continued, unab- 
ated, for five years, a decision was made 
to place an impermeable seal over the en- 
tire pile to restrict oxygen and water pen- 
etration into the fractured pyritic slate. 
As sufficient quantities of a relatively 
impermeable clay were present adjacent to 



52 



the pile, a compacted clay cap was chosen 
as the seal. Because Transport Canada had 
little experience in the mitigation of aci- 
dic drainage, an interdepartmental commit- 
tee was struck to advise on the "state-of- 
the-art" technology in this area. The Com- 
mittee charged the Canada Centre for Miner- 
al and Energy Technology (CANMET) to inves- 
tigate the feasibility of incorporating a 
salt layer between the pyritic slate waste 
and the overlying clay cap in order to 
reduce acid formation through bacteriosta- 
tic activity as well as through the oxygen 
depletion achieved by the clay capping. 
This paper describes the laboratory study 
that tested this hypothesis, the applica- 
tion of the salt-supplemented clay cap in 
the field and early indications of the ef- 
fects of the capping on acid drainage from 
the waste rock pile. 



Table 1 — Metals found in Halifax slates 



Element 



Percentage 



Aluminum 


9-13 


Iron 


3-6 


Manganese 


0.01-0.04 


Cobalt 




Nickel 




Lead 




Arsenic 




Copper 




Zinc 





ppm 



10-80 
40-270 
10-48 
4-19 
18-46 
70-210 



MATERIALS AND METHODS 



Table 2 — Clay overburden - percent pas- 
sing standard sieves (by weight). 



Laboratory 

Two laboratory studies were conducted 
to determine: (1) the effect of road salt 
on the production of ferric iron in a col- 
umn and (2) the effectiveness of a combina- 
tion salt/clay cap on the percolation 
rate. These studies were, in part, an out- 
growth of past work on the reclamation of 
solonetzic (sodic clay-pan) soils, where it 
was observed that Thiobacilli were unable 
to sustain viable populations in soils con- 
taining 4.8 meq of Na + /100 g of soil 
(McCready 1982, McCready and Krouse 1982). 
Later studies on the adaptation of Thio- 
bacilli to growth in saline waters indica- 
ted that T_;_ ferrooxidans could not tolerate 
sodium chloride concentrations greater than 
1% weight/volume. Thus, this laboratory 
study was conducted to test the hypothesis 
that the incorporation of a salt layer bet- 
ween the pyritic slate waste and the over- 
laying clay cap would provide bacteriosta- 
tic activity to reduce the formation of 
acid in addition to the oxygen depletion 
effect of the clay capping. 

Samples of Halifax slate and clay 
overburden from the Halifax Airport were 
provided to CANMET to carry out the labora- 
tory study. The Halifax slate is comprised 
mainly of slate and metagreywacke with 
minor quartzite. It is rich in sulphides, 
with pyrite and arsenopyrite and minor 
amounts of pyrrhotite, sphalerite and chal- 
copyrite. Whole rock laboratory analyses 
reveal metal contents in the ranges indica- 
ted in table 1. (Lund 1987.) 

A summary of the grain size analysis 
of four samples of the clay overburden is 
found in table 2. 



Sample 


19 mm 


9. 5 mm 


4.75 mm 


425 

micro- 
meter 


75 
micro- 
meter 


1 


100 


100 


97 


87 


63 


2 


100 


99 


96 


82 


53 


3 


100 


100 


99 


86 


60 


4 


100 


99 


98 


87 


65 



Effect of road salt on 
ferric iron production 

For the first laboratory study four 
columns ( 6 cm in diameter and 50 cm in 
length) were loaded with one kg of Halifax 
slate and overburden as shown in Fig. 1. 
In two of the four columns, 5 grams (less 
than 10 mm thickness) of road salt (NaCl) 
was placed on the surface of the slate. 
One salt-supplemented column and one unsup- 
plemented slate column were inoculated with 
500 mL of a 96-hour culture T\_ ferrooxidans 
grown on the low phosphate medium (McCready 
et al. 1986). The other salt-supplemented 
column and the remaining unsupplemented 
column were trickle leached with distilled 
water. The distilled water and the T. 
ferrooxidans cultures were circulated 
through the columns using a Gilson 
Minipuls-2 peristalic pump at a flow rate 
of 300 mL/h. Samples of the recirculating 
solutions were collected periodically from 
the two columns for pH determinations and 
ferric iron analyses. The total volume of 
solution for each column was kept constant 
at 500 mL by the periodic addition of dis- 
tilled water. 



A hydrometer analysis on sample 4 
revealed that 34% of the weight of the sam- 
ple is clay size particles (0.005 mm grain 
size or smaller). This overburden is des- 
cribed by the Unified Soil Classification 
System (U.S.C.S.) as Silty, Sandy Clay and 
Sandy, Silty Clay. (CL) . Its permeability 
is 4 X 10~ 7 cm/sec. 



Effect of a salt and clay capping 

In the second laboratory study, which 
assessed the effect of a salt (NaCl) and 
clay cap, two additional columns were pre- 
pared as shown in fig. 2. The test column 
had a layer of road salt placed between the 
pyritic slate sample and the clay layer 



53 



which was tightly compacted in both col- 
umns. Two hundred and fifty mL of distil- 
led water was added to each column to as- 
sess the percolation rate through the clay. 




Sample 

Glass wool 
Glass beads 
Glass wool 

Solution 



Figure 1. — Column design for testing the 
effect of 1.5% NaCl in solution. 




Glass wool 
Glass beads 
Glass wool 



Field 

Field testing of the salt-supplemented 
clay cap was carried out on the pyritic 
slate waste rock pile at Halifax 
International Airport, which, being 
constructed on a slope, ranged in depth 
from about 2 m to 10 m. The waste slate 
was surrounded on three sides by clay berms 
and was covered by a loose clay layer 
varying in thickness from 100 mm to 500 
mm. The top of the pile, although 
generally flat, had several depressions 
allowing water to pond. Side slopes varied 
from 2.5:1 to 4:1, which had led to 
significant erosion of the clay slope 
cover. The acid drainage collection pond 
is located at the base of the slope on the 
west side of the pile (fig. 3). 

The field work described here resulted 
from a contract awarded in October 1986 to 
construct surface drainage intercept chan- 
nels at the base of the north, east and 
south sides of the waste rock pile; reshape 
the top surface of the pile to provide a 
uniform gradient of 1.5%; supply highway 
coarse salt over the entire surface of the 
pile (including side slopes) at the rate of 
11.5 kg/m ; place a compacted clay cover 
750 mm thick over the entire area; supply 
and place 150 mm topsoil over this; place 
sod on the side slopes and areas of higher 
water flow on top; and hydraulically seed 
the remainder of the site. Fig. 4 is a 
typical section through the finished work. 

Work commenced immediately and by 
December 1986, when work stopped due to 
winter conditions, the top of the pile had 
been regraded, the coarse salt had been ap- 
plied, and a 375-mm thick compacted clay 
layer had been installed over the entire 
top surface. The surface water intercept 
channels were constructed in March 1987. 
Due to the low permeability and resulting 
high water retention of the salt-supplemen- 
ted clay layer, work was unable to start 
again until July 1987. The final 375-mm 
compacted clay layer was completed by early 
September, and the 150 mm of topsoil was 
placed by the end of September. Sodding 
was completed by October 15, 1986, but the 
hydraulic seeding remains to be done. 



Figure 2. — Column design for testing 
clay-capping. 



54 




hannels 



Figure 3. — Site plan. 



Sodding or Seeding 

Topsoil ( &f ' /■ ff'ffp 



i .- n i i H i i I . [ i I i 



Compacted Clay 

Highway Salt 
Slate 



'//s^/S^/A 



150 mm 

750 mm 
10 mm 



Figure 4. — Typical section through fin 
ished work. 

1.5 



RESULTS AND DISCUSSION 

Laboratory 

Effect of road salt on 
ferric iron production 

When the cultures of T. ferrooxidans 
were recirculated through the pyritic slate 
columns the soluble ferric iron content of 
the solution decreased due to adsorption of 
the iron to the clays in the slate/overbur- 
den mixture. After one month, the soluble 
ferric iron concentration of the leachate 
increased in the non salted column due to 
microbial activity (fig. 5). This was con- 
firmed by microscopic examination of the 
leachate; after 50 days the leachate con- 
tained about 10 organisms/mL. 



E 1.0- 



E 



0.5 




10 



20 30 40 

TIME (days) — 



50 



60 



Figure 5. -- Fe + ^ in test column effluents 



55 



In contrast, in the salt-treated col- 
umns, the soluble ferric iron content de- 
creased over the first 20 days and remained 
at a minimal concentration until the exper- 
iment was terminated. No active bacterial 
cells were observed during microscopic ex- 
amination of the leachate. 

In the two columns leached with dis- 
tilled water there was very little iron 
release and no T. ferrooxidans could be 
detected in the leachate. Extensive physi- 
cal breakdown of clay components of the 
overburden was observed in the salt- 
supplemented column. The continuous recy- 
cle of the saline solution most likely re- 
sulted in the stripping of calcium ions 
from the clays and the sodium ion replaced 
the calcium in the cation exchange sites. 
This resulted in the formation of clay 
fines which plugged the column after 35 
days. Repeated back-flushing of the column 
finally resulted in clearing of the clay 
fines and the experiment was continued to 



the 58 



th 



day, 



Effect of a salt and clay capping 

After compacting the clay and soil 
layers on the two columns, 250 mL of dis- 
tilled water was placed on the top of the 
columns. None of this water penetrated 
through the clay layer in either column. 
After two weeks the added water had evapo- 
rated and the soil and clay layers were al- 
lowed to dry out. 

The clay layer in the column without 
salt cracked vertically and on rewetting 
with 250 mL of water, the water penetrated 
through the clay layer and the column 
within 4 hours; the clay seal has not 
resealed to date. In contrast, the clay 
and salt seal in the other column cracked 
horizontally on drying. When 250 mL of 
water was reintroduced it took 12 hours for 
the water to penetrate through the clay and 
column. An addition of a second 250-mL 
aliquot of water, the clay resealed itself 
and no additional water percolation through 
the column has been observed to date. 



As stated above, the clay-capped 
column without salt has not resealed to 
date. After drying and cracking, with 
subsequent resettling of the clay, water 
percolation was observed to be 55 mL/day. 
After four months, the percolation rate has 
declined to 22 mL/day, and there has been 
0.75 increase in the effluent pH over this 
period (fig. 6). In contrast, the column 
with salt added sealed completely after 
drying and the subsequent addition of 500 
mL of water; to date no further effluent 
has been collected. Removal of an 
interstitital water sample indicated that 
the water had a pH of 2.98 with a very low 
iron content (48 ppm). 




I 2 3 
TIME ( months) 



Figure 6. — Volume and pH of the effluent 
from the clay-capped column without 
the salt layer 



The water which passed through these 
two columns initially was analyzed for iron 
content and pH . These values are presented 
in table 3. 

Table 3 — pH and ferric iron concentration 
of the initial effluent and after 250 
mL of effluent had passed through the 
clay sealed columns. 



Seal 



Effluent 



PH 



Fe 



+ 3 



ppm 



Clay alone First 5 mL 2.47 292 

After 250 mL 2.60 82 

Clay + NaCl First 5 mL 2.41 98 

After 250 mL 2.51 34 



Field 

As indicated in figure 3, all seepage 
from the waste rock pile and from the pave- 
ment subdrains beneath Taxiway D is collec- 
ted jointly in the treatment plant collec- 
tion pond west of the waste rock pile. The 
taxiway subdrain flows are directed to this 
pond by a single channel. Because flows 
from the waste rock pile enter the collec- 
tion pond along a seepage front, there are 
no specific quality nor quantity data 
available for this seepage. The quantity 
of seepage from the waste rock pile must 
therefore be calculated by subtracting the 
known taxiway subdrain flow volume from the 
known volume pumped through the treatment 
plant, while adjusting for changes in the 
level of the collection pond (fig. 7). The 
quality of the water at the plant intake 
represents a combination of the quality of 
water from the subdrains and from the waste 
rock pile. 



56 



Waste rock pile 

(unknown) 



Bypass 
(eliminate) 



? I 



Taxivay D 
subdrains 
(weirs in manholes) 



Volume change •— \ s~~\ 

(elevation change x surface area) V_/ \*^ 



Treaced 
(pump hours 
e pump rate) 



X + Y - A+B + C 



Figure 7. — Method of calculating seepage 
volume. 

Figures 8, 9, and 10 plot the concen- 
trations of acidity, iron, and aluminum in 
seep water at the outlet of the Taxiway D 
subdrains and at the inlet to the treatment 
plant, beginning 10 months prior to start 
of improvements and continuing to present. 



o 1986 
• 1987 



AT PLANT INTAKE 




^^ 



AT TAXIWAY SUBDRAINS 



JAM. FEB. MAR. APR. fW. JUN. JUL. AUG. SEP. OCT. NOV. OEC. 

TIME (months) — - 

Figure 8. — Concentration of acidity in 
seep water. 



o 

£ 



5 



o 1986 
• 1987 



AT PLANT INTAKE 




JAN. FEB. H*K. APR. HAY. JUN. JUL. AUG. SEP. OCT. NOV. OEC. 

TIME (months) 

Figure 9. — Concentration of iron in seep 
water. 



986 
987 



AT PLANT INTAKE 




AT TAXIWAY SUBDRAINS 
-neb 




FEB. MAR. APR. MAT. JUN. JUL. AUG. SEP. OCT. NOV. DEC. 

TIME (months) — - 



Figure 10. — Concentration of aluminum in 
seep water. 

All three parameters exhibit the same 
trend. During initial reshaping of the 
waste rock pile in November, 1986, a large 
rainfall on areas of freshly exposed pyri- 
tic slate caused concentrations of acidity, 
iron, and aluminum at the plant intake to 
jump to levels as high as or higher than 
previously recorded at this station. These 
increased concentrations did not diminish 
to pre-construction levels until mid- 
February of 1987, about three months 
later. With the salt layer and half the 
thickness of the clay cap in place, the 
1987 spring runoff resulted in much lower 
concentrations of the three parameters at 
the plant intake than for the previous 
year. Since that time all parameters at 
this station have been at concentrations 6% 
to 35% lower than the same period of the 
previous year. Because the concentrations 
of these parameters have at the same time 
remained relatively constant at the outlet 
of the taxiway subdrains, the improvement 
in the quality of water seeping from the 
waste rock pile has been even more pro- 
nounced. 

Because of variations in amounts of 
precipitation among seasons and from year 
to year, the discharge volume/unit of rain- 
fall during the same time period was used 
to determine the impact of the salt supple- 
mented clay cap on the volumes of water 
seeping from the waste rock pile. The tot- 
al volume of seep water was calculated per 
Figure 6 and then divided by the total 
rainfall during the same time period to 
yield the unit discharge volume. Figure 11 
shows this unit discharge for specific 
rainfall events for which all necessary 
data are available, beginning in August 
1986 (when an hour meter was first instal- 
led on the pump at the treatment plant) and 



57 



continuing to present. Figure 11 also il- 
lustrates the monthly average unit dis- 
charge volumes from the waste rock pile 
over the same time span. 



ol986 
.1987 



SPECIFIC STORM EVENTS 



i- 1986 



ft 198b 
® 1987 



MONTHLY AVERAGE 



1 
- 1987 2 






i! 



JAN. FEB. MAR. APR. MAr. JUN. JUL. AUG. SER OCT. NOV: DEC. 

TIME (months) 



Figure 11. — Waste rock pile 
discharge quantities. 



unit 



the pyritic slate and the clay cap resulted 
in a higher water retention time by the 
clay cap as indicated by the longer time 
required for drying and cracking of the 
clay layer. On drying, the clay cap crack- 
ed horizontally, rather than vertically, 
thereby preventing oxygen penetration to 
the underlaying slate. On rewetting, the 
salt induced swelling of the clay results 
in the reformation of an impermeable seal 
more quickly than in the absence of salt. 
In addition, if the clay should rupture, 
any water percolation would result in the 
leaching of soluble salt into the slate, 
thereby inhibiting any Thiobacilli within 
the water-penetrated slate. 

The salt-supplemented clay cover 
installed in the field on the waste rock 
disposal site has caused an initial de- 
crease in levels of acidity and dissolved 
iron and aluminum in the water seeping from 
the pile and a gradual decrease in the vol- 
ume of seepage from the pile. After one 
winter with only one half of the 
thickness of the clay in place, the effec- 
tiveness of the cap has not diminished. 
Monitoring is being continued to determine 
the long term effectiveness of this sealing 
method. 



Prior to start of construction of the 
cap, the unit discharge volumes over a 24- 
hour period for five specific rainfall 
events for which data are available lay 
consistently between 33 and 40 m-* of dis- 
charge/mm of rainfall. The storm during 
November 1986, on the freshly regraded 
waste rock pile resulted in a unit dis- 
charge volume from the pile of over 50 m^/ 
mm. 

Because of winter conditions, no more 
data are available until the end of April 
1987, at which time the salt layer and half 
the thickness of the clay cap had been 
installed. Unit discharge volumes for four 
specific storms varied from about 20 m-^/mm 
at this time to near zero in mid-August 
(when the clay cap was nearly completed). 
During three storms during late August and 
early September, after one of the driest 
summers on record, the unit discharge vol- 
ume rose inexplicably to 35 and then to 
over 50 m /mm. In six storms since this, 
the yield dropped to about 20 m^/mm or 
less . 



Literature Cited 

Lund, O. P. 1987. Acid Drainage from Min- 
eralized Slate at Halifax Airport. 
Proceedings Acid Mine Drainage 
Seminar/Workshop, Halifax, N. S. 
March 1987: 137-163. Environment 
Canada, Ottawa, Ontario. 

McCready, R. G. L. 1982. Bacterial oxida- 
tion of sulfur as a means of reclaim- 
ing Solonetzic soil. pp. 13.31. In 
J. C. Hermans ( ed ) , Solonetzic Soils 
in Alberta, Alberta Department of 
Agriculture, Edmondon, Alberta. 

McCready, R. G. L. and H. R. Krouse 1982. 
Sulfur isotope fractionation during 
the oxidation of elemental sulfur by 
Thiobacilli in a Solonetzic soil. Can 
J. Soil Sci. 62: 105-110. 

McCready, R. G. L. et al. 1986. Nutrient 
requirements for the in-place leaching 
of uranium by Thiobacillus f errooxi- 
dans. Hydrometallurgy 17: 61-71. 



The monthly average unit discharge 
volumes have trended downward from a high 
of 160 in October 1986 (just prior to start 
of improvements) to less than 40 in July of 
1987. Again, August of 1987 has an unex- 
plainably higher yield, although it is 
still lower than August of 1986. This 
yield held approximately constant through 
September and October, 1987. 

Conclusions 

In the laboratory study, preliminary 
work confirmed that road salt inhibited the 
production of ferric iron. In further 
work, the addition of a salt layer between 



58 



THE PARTITIONING OF FLOW COMPONENTS OF ACIDIC SEEPS FROM SURFACE COAL MINES 
AND THE IDENTIFICATION OF ACID-PRODUCING HORIZONS WITHIN THE BACKFILL 1 



Daniel T. Snyder and Frank T. Caruccio 2 



Abstract. — Two surface coal mine backfills in 
Upshur County, W.V. were monitored for 3 years to 
determine the relationship between the hydrology and 
the mine drainage quality. The recharge of the 
ground water system was determined by analyzing the 
response of the water table to precipitation events. 
Recharge occurs in two ways either by rapid recharge 
by runoff draining into highly permeable randomly 
oriented channels during times of high intensity - 
short duration rain events or regional downward 
migrating wetting fronts occurring during low 
intensity - long duration rains and spring thawing 
events. In the former case, water movement through 
the unsaturated zone was found to be from 12 to 30 
ft/day. Seep flow is comprised of shallow 
subsurface lateral flow (interflow) and deep ground 
water discharge (basef low) . Using regression 
analyses, water table elevations were correlated 
with baseflow and subsequently used to estimate the 
baseflow component of the total flow of the acidic 
seeps. Accordingly, total flow minus baseflow 
represents the interflow component of the seep. 
With these flow relationships, variations in seep 
chemistry were correlated with various flow 
components to identify the sources of acidity within 
the backfill. At one mine, interflow did not 
correlate with acid concentration, whereas at the 
second mine, interflow was negatively correlated 
with acidity. Therefore, runoff draining into the 
permeable channels of the vadose zone, flowing as 
interflow in the backfill, does not appear to be a 
major acid contributor. Instead most of the acidity 
is produced by low-intensity long-duration generated 
wetting fronts (spring recharge events) , 
infiltrating the system at field capacity. This 
creates a reservoir of acidic water that sustains 
the acid mine drainage seeps for the remainder of 
the water year. 



1 Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and Reclamation 
and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement) , April 
17-22. 1988. Pittsburgh, PA. 



2 Daniel T. Snyder is a Research 
Assistant and Frank T. Caruccio is 
Professor of Geological Sciences, 
Department of Geological Sciences, 
university of South Carolina, Columbia, 
SC. 



59 



INTRODUCTION 

Water is a crucial factor in the 
production of acid mine drainage, not only 
in the oxidation of pyrite but as a 
medium to transport the reaction products, 
including acidity and sulfate, to a 
discharge point. Previous studies of the 
relationship between ground water flow in 
backfills and acid mine drainage quality 
include Collier et al. (1970), Oertel and 
Hood (1983), Ladwig and Campion (1985), 
Lusardi and Erickson (1985) , and Erickson 
and Ladwig (1986) . The current study 
differs from the above studies in that it 
was designed to delineate the interflow 
and baseflow contributions to seep 
discharge, determine the relationship 
between the location of the flow path in 
the backfill and the resultant seep 
quality, and identify the acid-producing 
horizons within the backfill. This 
information can provide the basis for the 
design and development of in-situ acid 
mine drainage abatement strategies. 

METHODOLOGY 

Data Collection 

Two reclaimed surface coal mines in 
Upshur County, W.V. were selected for 
detailed hydrogeochemical analysis (fig. 
1) . The Eight Acre site is a small, 
unlined, northwestern portion of a larger 
15 ha refuse pile which has been almost 
completely covered with a 20 mil PVC 
plastic liner. A hiqh-volume acidic seep 
emerges from r.ne nortnwestern toe of the 
Eight Acre site. The Forty Acre Well site 
is a northwest-southeast oriented strip 
bounded to the north by a highwall. Acidic 
drainages are found at the western edge of 
this site. 

Beginning in the spring of 1985, 
alkaline recharge trenches were installed 
on the Forty Acre Well site to mitigate 
acid seeps through alkalinity additions 
(Caruccio, et al., 1984). Treatments were 
through the construction of permeable 
trenches that contained sodium carbonate 
briquettes and crushed limestone (designed 
to intercept and divert chemically 
modified surface runoff into the 
backfill) , and covering the surface of the 
mine with a veneer of crushed limestone. 

Thirteen observation wells were 
installed on the two sites between 1983 
and 1985 (10 on the Forty Acre site and 3 
on the Eight Acre site) . The 4 and 6- 
inch (10-and 15-cm) diameter cased (but 
not grouted) wells fully penetrate the 
spoil and are screened along a 10 ft (3-m) 
interval at the bottom of the wells. 
Water levels in these wells were measured 
on a weekly basis at the Forty Acre well 
site and daily at the Eight Acre site. 

During the time frame of this study 
seep discharge was measured daily at seeps 
6 and 7A which emanate from the Eight Acre 



site and the Forty Acre Well site, 
respectively. Flows are routed through 1- 
foot long, 8-inch-diameter, PVC pipes 
located as close to the seeps as possible. 
Discharge is measured using an 8-inch- (20- 
cm) diameter V-notch weir. 

Precipitation has been monitored on a 
daily basis at the site since 1981, using 
a 4-inch (10-cm) diameter all-weather rain 
gauge. 

Water samples have been collected from 
seeps 6 and 7A on a weekly basis since 
1983 and 1985, respectively. These 
samples are analyzed for temperature, pH, 
specific conductance, acidity, alkalinity, 
and sulfate at the Environmental 
Hydrogeology Laboratory at the University 
of South Carolina. 

Data Analysis 

Seep flow is comprised of shallow 
subsurface flow (interflow) and deep 
subsurface flow (baseflow) . Interflow is 
further defined as the ground water flow 
that occurs in the unsaturated zone, while 
baseflow is the ground water flowing in 
the saturated zone below the water table. 

To separate the total seep flow into 
its constituent components of baseflow and 
interflow, it was necessary to predict the 
baseflow components of the total flow. 
Using Darcy's law and relating the seep 
discharge to a function of the head, the 
baseflow component of the total seep 
discharge is directly proportional to the 
water table elevation measured in a well. 

At the Eight Acre site, the 
partitioning of seep 6 into its flow 
components was accomplished through 
multiple linear regression, to examine the 
relationship between the discharge of seep 
6 (during periods of little or no 
interflow input) and the water table 
elevation, as measured in wells 6, 6A, and 
6B. 

During the calibration of the 
regression model, discrete sets of 
observations were used to remove the 
influence that interflow might exert on 
the seep flow. Seep flow data were 
discarded if they were measured within 6 
days of a precipitation event of 0.25 inch 
(0.64 cm) or greater. Additionally, 
because snowmelt might contribute 
interflow for a much longer period after 
the precipitation occurred, all 
observations were discarded when 2 inches 
(5 cm) or more of snow was on the ground. 

Once calibrated, the regression model 
can be used to predict the baseflow 
discharge by substituting the estimated 
regression coefficients and intercept (as 
determined during the calibration process) 
and using the water table elevations 
measured in the wells. 

The seep flow from the Forty Acre 
Well site was analyzed in a fashion 



60 



MAINTENANCE AREA 




UNMINED 
SEEP 7 A 

FORTY ACRE WELL SITE 



GROUNDWATER DIVERSION 



LINER SITE 
COVERED WITH 20-md PVC 



Upshur County 
West Virginia 



LEGEND 



j° 


ACID SEEPS 








• 


MONrTORINC WELLS 








— — 2100 — 


_ STRUCTURAL CONTOURS OF THE COAL SEAM 


(feet) 






1 , 


500 
. . . 1 . . . 


1000 
..J 








SCALE (feet) 








Figure 1. Map of mine sites used in this study and their location. 



similar to that used for the Eight Acre 
site. Because the water levels in the 
wells at the Forty Acre Site are highly 
correlated with each other (see fig. 2), a 
single well was selected for use in the 
regression analysis of water elevation and 
seep discharge. Well 14 was selected 
because it had the highest r-square (r 2 ) 
value of any well when used as the 
regressor variable. 

Seep flow and water level observations 
used to calibrate the Forty Acre Well site 
regression model were selected using the 
same criteria as those used for the Eight 
Acre Site. 

Once the baseflow contribution to the 
total seep flow has been determined, and 
assuming that other contributions to seep 
flow, specifically overland flow and 
direct input from precipitation, are 
negligible, subtracting the predicted 
baseflow from the observed total seep flow 
yields the interflow. 



Having separated seep flow into its 
baseflow and interflow components, seep 
chemistry can now be correlated with the 
magnitude of each flow component to 
determine sources of acidity in the 
backfill. However, an important 
consideration should be recognized in this 
type of analysis. The total flow of the 
seep is an aggregate of ground water flow 
components whose origins are spatially and 
temporally distributed throughout the 
backfill. Correlation normally only 
evaluates the relationship between two 
variables without regard for temporal 
separation. While the effects of some 
variables are immediate, such as the 
transfer of the pressure head through the 
saturated zone due to a change in the 
elevation of the water table, the effects 
of other variables are delayed, such as 
the mass transfer of ground water from its 
place of infiltration through the mine- 
spoil to the seep. A time series analysis 
called cross-correlation, which creates a 
series of correlations comparing the 
effect of one variable with the delayed 



61 



impact on another variable (the unit of 
delay time is called the lag or lag time) , 
was used to circumvent this problem. 
Correlations are conventionally calculated 
from to n/4 lags, where n is the total 
number of lags in the time series (Davis 
1986) . 

RESULTS 

Water Table Response to Precipitation 

By comparing the well hydrographs with 
the precipitation data, the response time 
between a precipitation event and water 
table recharge in the vicinity. of a well 
can be established. In addition, due to 
the excellent data coverage for the wells 
at the Eight Acre site, cross-correlation 
can be used to determine the response time 
between precipitation and changes in water 
level elevations. 

The hydrograph of well 6 (fig. 3) 
shows the water table responding to 
precipitation events of about 1 inch or 
greater with a moderate increase in water 
elevation. Cross-correlation analysis 
indicates that well 6 responds in as 
little as 1 day and as long as 5 days, 
with a mean of 2 days. The unsaturated 
thickness of the spoil at well 6 is 
approximately 60 ft (18 m) . Therefore, the 
water that recharges the water table moves 
with an average downward velocity of 30 
ft/day (9 m/day) through the backfill. 

The hydrograph of well 6A (fig. 3) 
reveals a dramatic change in the response 
character of the well which occurred 
around March 1985, at which time post- 
holes spiked with caustic soda (NaOH) were 
constructed at the Eight Acre site to 
intercept flow channels, that feed Seep 
6A. Prior to March 1985, well 6A 
responded to precipitation in a manner 
similar to that of well 6 though with a 
slightly greater intensity and a faster 
response time. After March 1985 well 6A 
exhibited a flashy character, responding 
immediately and strongly to precipitation 
events of 1 inch (2.5 cm) or greater. 
Statistical analysis indicates that well 
6A has a minimum response time of less 

than 1 day and a maximum response time of 
1 day with a mean of less than 0.5 day. 
The unsaturated thickness of the spoil at 
well 6A is about 15 ft (4.6 m). Water 
recharging this area has an average 
downward velocity of greater than 30 
ft/day (9 m/day) through the backfill. 

The results from cross-correlation 
analysis of precipitation and the water 
elevation in well 6B show a consistent 
response time of 5 days. The mine-spoil 
unsaturated thickness at well 6B is 
approximately 90 ft (27 m) . If water is 
recharging vertically through the backfill 
in the vicinity of well 6B, then it has a 
downward velocity of 18 ft/day (5.5 
m/day) . However, the mine-spoil around 
well 6B is at the edge of a plastic liner 
installed to inhibit infiltration. There 



are several possible causes for the change 
in the water table as observed at well 6B: 
1) the liner may be leaking, 2) 
channelized flow paths may direct the 
water laterally through the unsaturated 
zone from the unlined portions of the 
mine, 3) ground water may be moving 
through the saturated zone downslope (with 
respect to the mine-floor) or a 
combination of (2) and (3) . 

At the Forty Acre Well site all wells 
display similar responses to precipitation 
events, albeit with varying magnitudes 
(fig. 2) . Response times appear to be on 
the order of a few days. 

Collier et al. (1970) have also 
observed well responses to precipitation 
events within 24 hours at a mine-spoil in 
Beaver Creek Basin, KY indicating that 
recharge is capable of moving rapidly 
through the unsaturated zone. 

Relationship Between Baseflow 
and Water Table 

Using regression analysis, a strong 
predictive relationship (adjusted r 2 = 
0.89 and a p-value (or probability) = 
0.001) was established between baseflow 
and water levels in the observation wells 
at the Eight Acre site. The baseflow 
hydrograph predicted for seep 6 and the 

total seep discharge hydrograph compare 
favorably (fig. 4) . Subtracting the 
baseflow from the total seep discharge 
yields the predicted interflow hydrograph 
which is also presented in figure 4. 
Significant interflow events appear to be 
correlated 

with the precipitation record, suggesting 
that the regression model is working 
properly. 

At the Forty Acre Well site an equally 
strong predictive relationship (adjusted 
r 2 = 0.89, p-value = 0.001) exists between 
baseflow and the water level in 
observation well 14. 

For the period following April 1985, 
at which time induced alkaline recharge 
zones were introduced at the Forty Acre 
site, the predicted baseflow hydrograph 
for seep 7A agrees favorably with the 
total seep discharge (fig. 5) . Prior to 
April 1985 the predicted baseflow is 
consistently greater than the observed 
total discharge. The model was calibrated 
using data following the treatment, 
suggesting that the hydrology of the site 
has been substantially modified by the 
reclamation efforts. 

Flow Components and Seep Chemistry 

The cross-correlation analysis of 
baseflow and acidity for the Eight Acre 
site shows that at zero lag time acidity 
is negatively correlated (-0.57) with 
baseflow. The correlation becomes 
progressively more positive with 
increasing lag time, reaching a peak 



62 



2047 

2046- 

2045 

2044- 

2043- 

2042 

2041 

2040 

2039 

2038 

2037 



FORTY ACRE WELL SITE WELLS: WATER TABLE vs. TIME 




PRE-TREATMENT POST-TREATMENT 



.l.Lll^liiLllL I. ,.|llllilll I J.Jl |1 .j|j||| f J|.L [ Jl^l|.ll lll, f L.Hli. 
FMAMJJASOND ' F M A M J J A ! 



Jill 



Nd'fMAMJJASOND'FM 



#, 



I..H. ■!■ 



.it k ■! 



m-r 



-3 =J 

O 

-2 2 



1987 



1984 1985 1986 

Figure 2. Water table elevations versus time and precipitation data for Forty Acre Well site. 



2069 

2068 -i 

2067 

2066 

2065- 

2064- 

2063- 

2062- 

2061 - 

2060 - 

2059 



EIGHT ACRE SITE WELLS: WATER TABLE vs. TIME 




PRE-TREATMENT POST-TREATMENT 



Ljklw i i ..lli.iJn LijLikl f .iJ|.>.. [ JiL.iJ..J 1 .iii. f .L 

F M A MJJASOND I FMAMJJ 



"vT-J 



1984 



S N ' FMAMJJ 

1985 



ftt 



-WELL 6B 




iiUjillljLut .I.,, , liLk.ili 



N D ' FMA M 3 J A S N~ 
1986 



i 

3 =1 

O 

H2 2 

1 "? 



1987 



o i 



Figure 3. Water table elevations versus time and precipitation data for Eight Acre Well site. 



63 



correlation of +0.60 with a lag time of 20 
weeks. 

The initial inverse relationship may 
be the result of three effects. The first 
is that an increase in baseflow is usually 
preceded by an increase in interflow. As 
described below, interflow is negatively 
correlated with acidity and may mask the 
impact of baseflow. A second alternative 
is that the baseflow, which arrives at the 
seep earliest, is from meteoric waters 
which have infiltrated the backfill in the 
vicinity of the seep. The thickness of 
the mine-spoil through which these waters 
pass is relatively thin as compared with 
most of the Eight Acre site. 
Additionally, this particular part of the 
mine-spoil is adjacent to the outslope, 
which has more oxygen available to 
circulate and react with the acid- 
producing material. Therefore, the 
infiltrating water comes into contact with 
a thinner sequence of overburden material 
that may already be leached of its acid 
production potential. The third 
possibility is that the installation of 
post-holes treated with sodium hydroxide, 
centered around the seep, is reducing the 
acid production in the vicinity of the 
seep or neutralizing acidity. 

At the Eight Acre site the strong 
positive correlation between b.aseflow and 
acidity which peaks at a lag of 20 weeks 
may represent the arrival of water that 
has percolated through the spoil up- 
gradient from the seep and flowed 
laterally below the water table. In this 
area the spoil is thicker and the oxygen 
circulation may be more restricted than in 
the spoil near the seep. These factors 
could prolong the acid-producing potential 
of this portion of the backfill and 
increase the acidity of the water relative 
to the water that passes through a thinner 
portion of the spoil. 

A weak negative correlation (-0.26) 
exists between interflow and acidity at 
zero lag time. Though the correlation 
becomes slightly more negative (-0.33) 
over the next two lag periods, interflow 
is expected to have diminished by the end 
of the first week. The results show that 
interflow serves to dilute seep acidity. 

Acidity to sulfate ratios, when 
compared with baseflow, indicate slightly 
higher acidity to sulfate values with 
increasing lag time. This may be due to a 
difference in the pyrite oxidation 
reactions. 

Acidity to sulfate ratios display an 
inverse relationship with interflow (-0.43 
at 1 week) . This may indicate that some 

neutralization of acidity is occurring, 
possibly due to treatment interaction. 

At the Forty Acre Well site, cross- 
correlation analysis shows a very weak 
relationship between baseflow and acidity. 
Initially, the correlation is negative 



(-0.19 at zero lag time). After 3 weeks 
there is a slightly positive correlation 
(+0.20), and at 11 weeks there is another 
negative peak (-0.24). The early negative 
correlation may be the result of causes 
similar to those of the Eight Acre site. 
The overall subdued character of the 
relationship between baseflow and acidity 
may be due either to the character of the 
ground water flow within the mine site or 
to treatment effects which will be 
discussed below. 

There is no apparent correlation 
between interflow and acidity from zero 
lag time up to 20 lags (weeks) . 

At zero lag time, sulfate is 
negatively correlated (-0.43) with 
baseflow. The correlation becomes more 
positive and reaches a peak (+0.47) near a 
lag of 16 weeks. This is similar to what 
was observed at the Eight Acre site. Here 
too, the local baseflow may pass through a 
thin column of spoil which may be depleted 
of its acid production potential. The 
baseflow originating from more distant 
parts of the mine may have a higher 
acidity because of traversing a thicker 
and more reactive spoil. 

There is no apparent correlation 
between interflow and sulfate. This 
suggests that the observation that acidity 
is not correlated with interflow is due to 
the lack of acid production in the 
unsaturated zone through which the 
interflow moves, rather than 
neutralization of the acidity. 

The acidity to sulfate ratio exhibits 
a slight positive correlation (+0.29), 
with baseflow peaking at 3 weeks followed 
by a moderate negative correlation which 
reaches a peak at 13 weeks (-0.56). The 
inverse relationship may indicate 
neutralization of acidity in ground water 

within the backfill due to the impact of 
treatments or an alteration of the 
chemical reactions controlling the 
oxidation of pyrite and its reaction 
products . 

There is no apparent correlation 
between acidity to sulfate ratio and 
interflow. There is no evidence to 
suggest that the interflow water quality 
is being modified by neutralization or an 
alteration of the oxidation reactions of 
pyrite. 

CONCLUSIONS 

During high-intensity, short-duration 
precipitation events, runoff recharges the 
water table by flowing through highly 
permeable channels in the backfill. This 
occurs quickly as demonstrated by the 
rapid response of the water table to a 
precipitation event (manifested by the 
rise of the water elevation in observation 
wells) . Under these conditions, the rate 
of water movement through the unsaturated 
zone is from 10 to 30 ft/day (3 to 9 



64 



25000 -i 



EIGHT ACRE SITE SEEP 6: FLOW COMPONENTS vs. TIME 



20000 
15000 
10000- 
5000- 
0- 



UJ 

a 

a: 
< 

x 
o 
tn 
o 



15000 - 
10000 - 
5000- 
0- 
-5000- 
•10000- 




25000 
20000 
15000 
10000 
f-5000 






PRE-TREATMENT POST-TREATMENT 



,ii,|j,MiL|ii i v ,|iijIii L i^jiiiJA. j^iI^iilLim.. ., .1 Lil i.iii.ilhilLkJj. i-i, .i JlLi., ..I . .. 

fmAmjjAs6nd i fmA mjJas6ndIfmamJJaso 
1984 1985 1986 



OBSERVED TOTAL FLOW 




PREDICTED BASEFLOW 




PREDICTED INTERFLOW 



S N D 



— 5 

— 4 

— 3 

2 

— 1 
-0 



1987 



Figure 4. Flow hydrographs of total flow, predicted base flow 
and predicted interflow for Seep 6, Eight Acre Site. 



50000 
40000 - 
30000 ■ 
20000 - 
10000 ■ 




FORTY ACRE WELL SITE SEEP 7A: FLOW COMPONENTS vs. TIME 



Ul 

o 

I 
o 
en 



40000 - 
30000 - 
20000 - 
10000 - 
0- 
-10000- 



50000 
40000 
30000 
20000 
10000 




1984 




PRE-TREATUENT POST- TREATMENT 




, 1 1 .JJihIii 1 , i .JijJljl |U|iJjujjj|i [.iji... .. J ),ll ilJtliliJj, i..., ,i ,|lil ii„ .,L 

M J J AS OND I FMAMJJASOND ' FMAMJJ ASON 



OBSERVED TOTAL FLOW 




PREDICTED BASEFLOW 




PREDICTED INTERFLOW 



5 
— 4 

3 



— 1 




1985 



1986 



1987 



Figure 5. 



Flow hydrographs of total flow, predicted base flow 
and predicted interflow for Seep 7A, Forty Acre Site. 



65 



m/day) . The speed at which the water 
table is recharged reflects very high 
hydraulic conductivity and/or channelized 
flow. 

Further analysis of the data show the 
installation of induced alkaline recharge 
zones at both the Eight Acre and Forty 
Acre Well sites to have successfully 
modified the hydrology of the backfill. 

The plastic liner covering a major 
portion of the Eight Acre site is not 
completely effective in preventing 
recharge of the backfill by meteoric 
waters due either to unlined portions of 
the backfill, leaks in the liner, of 
lateral flow. 

Regression analysis was successfully 
used to correlate water table elevations 
with baseflow (sustained by deep ground 
water regimes) and to predict the baseflow 
component of the total 'flow. Having 
successfully estimated the baseflow 

component, it was possible to guantify the 
interflow (shallow lateral flow) component 
of a seep. 

The baseflow at both the Eight Acre 
and Forty Acre sites promptly displays a 
negative correlation with acidity. This 
indicates that the interflow component of 
baseflow, which arrives at a seep first, 
has less acidity because it flowed through 
a smaller thickness of spoil and which may 
be leached clean of its acid-producing 
potential. 

At both the Eight Acre and Forty Acre 
sites baseflow exhibits a delayed positive 
correlation with acidity. This suggests 
that baseflow, sustained by ground water 
from deeper parts within the mine-spoil 
has a higher acidity, possibly due to the 
greater thickness of spoil through which 
recharge must pass and/or a greater acid- 
producing potential of the spoil material. 
Under these conditions recharge to the 
backfill is by the occasional areally 
distributed, vertically infiltrating 
wetting fronts, generated by low- 
intensity, long-duration rains or spring 
thaws. Here, the entire backfill becomes 
saturated and a larger volume of rock 
material is leached. 

At the Forty Acre Site interflow did 
not correlate with acid concentration. At 
the Eight Acre site interflow is 
negatively correlated with acidity. 
Therefore, the vadose zone of the 
backfill, through which interflow moves, 
does not appear to be a major acid 



contributor. This suggests that the 
acidity of the seep is sustained by a 
large acid reservoir, occasionally 
replenished by long-term vertically 
infiltrating waters (recharge events 
occurring at field capacity saturation) 
that commonly develop during the spring 
thaw. 



LITERATURE CITED 

Caruccio, F. T. , G. Geidel, and R. 

Williams. 1984. Induced Alkaline 
Recharge Zones to Mitigate Acidic 
Seeps, p. 43-47. In Proceedings, 
Symposium on Surface Mining Hydrology, 
Sedimentology, and Reclamation. Univ. 
KY, Lexington, KY. 

Collier, C. R. , R. J. Pickering, and J. J. 
Musser, (Eds.). 1970. Influences of 
Strip Mining on the Hydrologic 
Environment of Parts of Beaver Creek 
Basin, KY, 1955-1966. 80 p. USGS 
Prof. Paper 427-C. 

Davis, J. C. . 1986. Statistics and Data 
Analysis in Geology. 646 p. John 
Wiley, New York. 

Erickson, P. M. and K. J. Ladwig. 1986. 
Field Observations of Potential Acid 
Sources within Surface Mine Backfills. 
9 p. (addendum) . In Proceedings, 
West Virginia Surface Mine Drainage 
Task Force Symposium, Morgantown, WV. 

Ladwig, K. J., and P. Campion. 1985. 

Spoil Water Quality Variations at Two 
Regraded Surface Mines in 
Pennsylvania. p. 121-130. In 
Proceedings, Symposium on Surface 
Mining Hydrology, Sedimentology, and 
Reclamation. Univ. KY, Lexington, 
KY. 

Lusardi, P. J., and P. M. Erickson. 1985. 
Assessment and Reclamation of an 
Abandoned Acid-Producing Strip Mine in 
Northern Clarion County, Pennsylvania, 
p. 313-321. In Proceedings, Symposium 
on Surface Mining Hydrology, 
Sedimentology, and Reclamation, Univ. 
KY, Lexington, KY. 

Oertel, A. O. , and W. C. Hood. 1983. 
Changes in Ground Water Quality 
Associated with Cast Overburden 
Material in Southwest Perry County, 
Illinois. p. 73-79. In Proceedings, 
Symposium on Surface Mining Hydrology, 
Sedimentology, and Reclamation. Univ. 
KY, Lexington, KY. 



66 



THE USE OF PHOSPHATE MATERIALS AS AMELIORANTS FOR ACID MINE DRAINAGE 



John J. Renton, Alfred H. Stiller and Thomas E. Rymer 



Abstract. — Research has shown that limestone, when used as 
an acid ameliorant in surface mine reclamation may result in an 
increase in both the acid production rate and the total acid 
load. Rock phosphate has been shown to be an effective acid 
ameliorant based upon its ability to sequester the major oxi- 
dizing agent of the iron disulphide minerals, Fe + ->, as an in- 
soluble phosphate. A series of coordinated bench scale and 
small field scale experiments were conducted to evaluate the 
response of rock phosphate intermixed with toxic rock materials 
upon exposure to laboratory and real weathering conditions. 
The experiments utilized rock phosphate from granule size down 
to -325 mesh at application rates up to 7 wt% apatite. The 
work showed that rock phosphate in excess of about 1/16" mean 
diameter was relatively ineffective as an ameliorant. Effect- 
iveness as an ameliorant, however, increased with decreasing 
particle size and application rate. Most effective was a clay 
slurry spiked with -325 mesh apatite and hydraullcally applied. 
The results showed that a minimum application rate of about 1 
wt% apatite is required for ameliorization to become effective 
and that reductions of acid production rate and total acid load 
can be reduced up to 90% over an untreated control at applica- 
tion rates of about 4 to 5 wt% apatite. 



INTRODUCTION 

In areas producing high sulfur coal (>1 wt% S) 
such as the Illinois Basin, highly calcareous over- 
burden materials or calcite-rich surf icial materials 
are readily available for reclamation purposes. As 
a result, acid mine drainage is rarely a problem. 
However, in high sulfur areas such as the northern 
Appalachian Basin where calcareous rocks are not 
generally abundant, limestone has been routinely 
added to the toxic (sulfur containing) rock mater- 
ials during reclamation to ameliorate acidity. How- 
ever, the use of limestone has not always resulted 
in the elimination of acid waters. 

Early in the research of the authors, experi- 
mental weathering data indicated that the presence 
of calcite in sulfur-containing rock materials 
could actually INCREASE acid production. In initial 
soxhlet leach experiments designed to evaluate the 
acid producing potential of various coals and coal- 
associated rocks, a direct correlation was observed 
between the calcium content and INCREASING sulfate 
concentration (used as a measure of originally 



produced acidity) of the leachates. It was also ob- 
served that coals collected for study which did not 
have exceptionally high sulfur contents but contain- 
ed iron disulphide minerals in direct contact with 
calcite invariably would rapidly decompose the cloth 
collection bags. Acid was obviously being produced 
at a rate disproportionate to the sulfur content of 
the coal. 

To further investigate the effect of calcite on 
acid production from toxic rock materials, a series 
of experiments was conducted which intermixed pure 
calcite with several rock materials of differing 
toxic potential at application schedules of 0.50, 
1.00, 2.00 and 4.00 wt% calcite. The experiments 
showed that the addition of calcite in application 
rates up to about 5 wt% increased acid production 
over that of the untreated controls (Fig. 1). Acid 
loads reached a maximum at about 1.0 wt% calcite 
addition after which acid production dropped. At 
about 5 wt% CaC03 addition, the acid production 
equaled that of the untreated controls. Beyond 5 
wt% calcite addition, the acid production decreased 
relative to the controls, and calcite became an 



67 




WI. % Calclto Addition 



FIGURE 1. Relation Bn.itn S0 4 " a Production (Acid Production) from 

Four Typical Toxic Rock Matartala Traatad with Calelta at 0.5. 
1.0. 2.0 and 4.0 wt. % Addition Rataa 



effective acid ameliorant. 

It is important to note that the use of lime- 
stone in surface mine reclamation is usually not in- 
tended to eliminate the PRODUCTION of acid from the 
weathering of rocks containing iron disulphide min- 
erals but rather is meant only to neutralize acid- 
ity. It should be pointed out, however, that 
CaC03 will inhibit the oxidation of the iron disul- 
phides and subsequent acid production if the appli- 
cation rate is high enough to raise the pH of the 
system to approximately pH 6-7. At this point, 
acid production would be reduced both by the precip- 
itation of dissolved iron, the major oxidizing 
agent, and by the inhibition of iron and sulfur 
oxidizing bacteria. 

Normally, calcite is used simply to neutralize 
acid already generated. Because the neutralization 
reaction takes place at the surface of the calcite 
grains, the calcite grains will eventually become 
coated with precipitated iron oxy-hydroxides there- 
by reducing the effective calcite concentration. 
According to the experimental data presented in 
Fig. 1, should the effective calcite concentration 
drop below 5 wt%, the system will once again become 
a net acid producer. This may explain the frequent 
lack of success in using limestone as an AMD amel- 
iorant. 

With limestone being questioned as a potential 
ameliorant, a substitute was sought. Our objective 
was to utilize a natural occurring material so as 
not to subject the environment to another source of 
unnatural pollution and preferably to utilize a ma- 
terial that would actually prevent the oxidation of 
the iron disulphide minerals rather than simply neu- 
tralize acidity. Apatite was chosen because it fill- 
ed all the criteria. Apatite eliminates acid for- 
mation by removing Fe + ^, the major oxidizer of the 
iron disulphide minerals, and precipitating it as 
an insoluble iron phosphate below pH 5.5. In addit- 
ion, apatite only dissolves when the conditions drop 
below pH. 5.5 (ref. 1). It is therefore a time re- 
lease material. The purpose of this work was to 
systematically evaluate the effectiveness of a 
variety of available phosphatic materials as acid 
ameliorants . 

EXPERIMENTAL MATERIALS 



phate slurry (slime) . The sand to pebble sized 
rock phosphate, composed largely of the mineral 
apatite, is the material that is presently being 
used in the Appalachian Basin with varying amounts 
of success (or lack of success) . The fine grained 
rock phosphate is the larger sized apatite that has 
been ground into fine particle sizes in order to 
increase the available reactive surface area. In 
this research, two sizes of ground apatite were 
used, one designated CODE 30 that was 95% in the 
size range from 150 to 250 mesh and a second mater- 
ial designated CODE 31 that was about 65-70% less 
than 325 mesh. The 150-250 mesh material was con- 
sidered the finest size that could be used on-site 
without dust generation problems. The -325 mesh 
material was primarily considered as an additive to 
increase the apatite content and subsequent effect- 
iveness of the dried slurry material. 

The phosphatic slurry (slime) is a reject ma- 
terial of the phosphate mining industry. The uti- 
lization of the dried slurry was considered most 
important in that it was an attempt to utilize a 
costly refuse material of the phosphate industry to 
solve an equally costly problem of the coal indus- 
try. The composition of slurry material varys de- 
pending upon the geographic source. The dried 
slurry used in these experiments averaged 25 wt% 
apatite the remainder being smectite dominated clay 
minerals. The highly reactive clay-sized apatite 
content, the ability of the dried material to be 
reslurried and hydraulically applied, the tendency 
of the slurry to stick to applied surfaces and the 
fact that the apatite content could be increased 
by "spiking" with fine-grained (-325 mesh) apatite 
were the main attributes of the material. 

TOXIC MATERIALS: The toxic materials used in this 
work were of two varieties. The first was a common 
toxic "standard" material, a typical coal cleaning 
plant waste which was used in both bench scale and 
small field scale experiments. In addition, all 
phosphatic materials were tested in bench scale 
experiments against a second suite of toxic mater- 
ials which represented all of the various rock 
lithotypes encountered in coal mining. The second 
suite of rock materials provided the range of sul- 
fur contents and acid producing potentials normally 
encountered in mining. The data generated by the 
experiments using this suite of toxic materials 
allowed the statistical testing of all possible in- 
terrelationships between the various phosphate 
addition rates and individual rock compositional 
and acid parameters. 

EXPERIMENTAL METHODS 

The experiments conducted in this work were of 
two types: 1) bench scale and 2) small field 
scale (barrel) experiments. The bench scale 
experiments exposed 100 gram samples of the toxic 
materials and toxic material/phosphatic material 
mixtures to the synthetic weathering conditions 
while the small field scale experiments exposed 300 
pounds of the experimental materials to natural 
weathering conditions. Both procedures were re- 
ported earlier and described in detail by the auth- 
ors (2,3). The experimental arrangements for each 
is shown in figures 2 and 3 respectively. 



PHOSPHATE MATERIALS: Phosphate materials were cate- 
gorized for this study into three types: 1) sand to 
pebble sized rock phosphate, 2) fine grained (less 
than 150 mesh) rock phosphate and 3) dried phos- 



68 



SOXHLET EXTRACTOR 



Reactor Section 



Extraction Thlmbfi 




Vapor By-Piat Tub* 
Syphon Tub* 



FIGURE 2. Schem.dc Diagram of Soihl.t Extractor 



TABLE 


1 FCTERI 


;e:.tal se 


? *1 












CC# 


COAL BED 


LITHIC 


St 


A1FHA 


s 3oo 


ETA 


TREATMENT 


:reatjx»t 




ASSOC. 


TXPE 










cc# 


6*34 




KEF. 


2.9T0 


-0.0003 


2.56 


46.00 


1/4--1/8" 


5192 


116 


K.KITT 


ROOP 
oKALE 


1.U10 


-0.CO12 


7.578 


22.415 


1/4--I/8" 


^192 


lift 


U.Ft'EE 


KEF. 


2.830 


-0.0093 


31-159 


25.287 


l/*"-l6" 


5192 


22 




REF. 


2.7*2 


-0.0012 


10.778 


12.*05 


i/**-i/a- 


5192 


6*3<. 


U.FPEE 


P-EF. 


2.970 


-0.C0OJ 


2.56 


*8.t)0 


1/3--1/16" 


5191 


122 


U.FHEE 


S.'iALE 


7.271 


rO.^075 


37.3*7 


23.S9* 


l/8"-l/l6" 


5191 


60 


O.FBEE 


2.955 


-O.P012 


7.925 


10.*6T 


1/8--V16- 


5191 


53 


u.?a:c 


pi?! 


6.50: 


-C.0C.6 


26.6J0 


25.371 


1/8" -1/16- 


5191 


6*3* 


U.F3EE 


RES'. 


2.970 


-r.ooo3 


2.56 


«e.oo 


<1/16" 


5190 


326 




REP. 


7.010 


-O.C039 


33-:£3 


22. jl* 


<1/16" 


5190 


35 


U.FPEE 


REP. 


1.1*0 


-0.0019 


9.957 


32.6*9 


<1/16- 


5190 


19 




REF. 


1.812 


-0.3010 


e.51* 


23.103 


<1/16" 


5190 


330 


U.FKEE 


EARTH 


?.6d0 


-0.0068 


20. M5 


6.263 


CODE 30 


5168 


*7 


M.K1TT 


5-263 


-0.0007 


19.544 


23-987 


CODE 30 


5168 






COAL 














65 


U.FP.EE 




6.33" 


-C.0011 


20.333 


20.158 


CODE 30 


5168 


DO 




?kP~zr-Q 


*.2*2 


-c.oois 


21.26 2 


30.1"3 


CODE 30 


5168 


6*3* 


•J.Ff.ZZ 


SEP. 


2.970 


-O.C003 


2.56 


*6.000 


CODE 30 


5168 


33T 




SEP. 


9. Hi 


-C.0051 


35.06* 


21.017 


CODE 30 


5168 


320 


U.F3EE 


RCF. 


1.535 


-C.0535 


5.795 


25.32* 


CODE 30 


516B 


123 


U.FhEE 


KEF. 


3.136 


-O.0OBO 


19-570 


I6.6fi0 


CODE 30 


5168 


121 


U.mEC 


ROOF 
SKALS 


9-7*0 


-0.DOK2 


13.737 


24.710 


CODE 30 


51'JH 


63 


u.rnns 


REP. 


*.56o 


-0.0J32 


2*. 22 


21.13 


CODE 30 


5163 


55 


U -FREE 


HOOP 
S!!ALC 


6.310 


-0.0006 


9-123 


27.71 


CODE 30 


5168 


2* 


J. FREE 


overs. 

SHALE 


0.652 


-0.0050 


*.83 


5.00 


CODE 30 


5163 


6*3* 


U.PSUE 


Hi'S. 


2.197 


-0.0003 


2.56 


46.J0O 


ACRICO 


5166 


5070 


U.FrtEB 


RBK. 


3.150 


-O.0J50 


17.06 


5d.36 


AG HI CO 


5166 


13 


U.TREE 


R«Si*. 


(.560 


-.1.0052 


16.53 


17.73 


A0R1C0 


5166 


12 


L.KX5S 


HOOF 

SHALE 


1.586 


-0.0117 


1*.6< 


5.37 


AQRICO 


5166 


19 






!.Cl2 


-0.0010 


*.777 


23.103 


ACRICO/C3I 


P3*8 


26 


U.FIiEt. 


HOOF 

Si, ALE 


2.428 


-0.0063 


20.702 


?j.676 


A3KIC0/C31 


83*8 


09 


U.KflEE 


s.s. 


1-7*0 


-0.0003 


1.359 


45.210 


AGHIC0/C31 


33*8 


62 


U.TREE 


HEP. 


5.525 


-J.TJ50 


43. 102 


20.9*3 


ACRIC0/C31 


83*8 


64 


L\?rEL 


REF. 


3.312 


-0.0061 


27.3*7 


15-911 


ACRIC0/C31 


i3*8 


119 


L.FSEE 


PA,-TI!tO 


1.9*0 


-0..1033 


12.10* 


12.114 


ACHIC0/C31 


83*8 



Packing Material 




FIGURE 3. Schamallc ol Small Fiald Scala (Dacral) Expa'lmenl 



BENCH SCALE EXPERIMENTS 

The bench scale experiments were designed to 
evaluate the effect of different addition schedules 
of each of the three basic phosphatic materials on 
both the rate of acid production and on the ultim- 
mate acid load. Four sets of experiments were con- 
ducted. 

SET #1: 

Table 1 summarizes the first set of bench 
scale experiments which were conducted to evaluate 
the effect of intermixing the various kinds of phos- 
phatic materials with a suite of different toxic 
lithotypes. In each of the experiments, the respec- 
tive phosphatic material was added in 0.25, 0.50, 
1.00 and 2.00 wt% apatite. All data were compared 
to an untreated control. All experiments were run 
in triplicate. The table lists the coal bed assoc- 
iation, the specific lithotype and the various acid 
evaluation parameters for each of the toxic mater- 
ials. The CC#'s are laboratory identification 
numbers . 



pebble sized rock phosphate separated into the in- 
dicated categories. As previously indicated, the 
"CODE 30" material is the rock phosphate in the 
size range from 150 to 250 mesh and the material 
indicated "CODE 31" is the same material ground to 
pass 325 mesh. The CODE 30 and CODE 31 materials 
were acquired from Texasgulf Sulfur at Aurora, 
North Carolina. The material termed "AGRICO" is a 
solar dried slurry (slime) acquired from the 
AGRICO Mining Co., Mulberry, Florida. 

The results of the experiments were compared 
by plotting percent acid reduction relative to the 
control versus the wt% apatite addition. The data 
were extrapolated by computer to 5 wt% apatite 
addition. 

SET #2: 

Based upon the success of the initial experi- 
ments, a second set of experiments was conducted 
utilizing the AGRICO slurry material with and with- 
out CODE 31 addition intermixed as a water slurry 
and run against a common toxic material. The ex- 
periments are summarized in Table 2. Each wt% apa- 
tite addition with the exception of 7.00 wt% apa- 
tite was conducted using two different slurry mix- 
tures. All experiments were conducted in tripli- 
cate. Based on the results of these experiments, 
two final sets of experiments were designed. 



TABLE 


2 EXPEKir-EH'TAL SET 




SOLID I1II 


CCf 


go AORICO 


£D CODE 31 


83*2 
83*3 
83i» 


5 

260 

260 



13 

13 


83»5 
83*6 


300 
300 


60 
60 


8347 


"50 


225 


83*8 


«50 


225 


83*9 


450 


225 



SLURRY 1 


IX 








ia SOLID 


ut; apat 






MIX 


ADDITION 






5 




0.25 










1. 00 






225 




1.25. 


1.75. 


1.85 






3.50. 


3-75. 


4.00 


60 




0.75. 


2.25. 


2.75 


300 




2.75. 
".50 


3.25. 


a. 00 


75 




0.25. 


0.50. 








1.25. 


1-50. 


1.75 






2.50 






150 




0.50. 


1.00. 


2.00 






2.50, 


3.00. 


i.io 


• 50 




0.75. 

3.00. 
5-00. 


1.50. 
3.75. 
7.00 


2.25 

4.50 



SET #3: 

In this set of experiments, slurry CC# 8348, 
a 1:0.5 blend of dried slurry CODE 31 apatite was 
intermixed with four different toxic materials at 
application schedules of 1, 2, 3 and A wt% apatite 
(see Table 3). 



The first 12 experiments listed are the sand to 



69 



1„ULE 3 EXPKRIMUITAL SET «3 

ACRICO VS. 6 DIl-'FSREtlT TOXIC MATERIALS 



CCt 

26 
62 
64 
119 



COAL 

U.FREE 
U.FRES 
U . KREE 
L.FREE 



ROCK TVP. 

ROOF SH. 
PREP.REF. 
PKEP.REF. 
PARTING 



St 

2.4 18 
5.525 
3.312 
1.940 



ALPHA 

0.0065 
0.0050 
0.0061 
0.0033 



S300 

21.325 
23.397 
17.741 
13.976 



ETA 

25.676 
20.94 3 
15.911 
12.114 



CC# 8348 SLURRY 

450 sm ACRICO + 225 bin CODE 31 
750 nl H 2 + 150 gra MIXTURE 

APPLICATION SCHEDULE 

CONTROL, 1 wt%, 2 wtX, 3 wt% and 4 wtj APATITE 

SET #4: 

Initially, the CODE 30 material was tested a- 
gainst a variety of toxic lithotypes. In the 
fourth set of experiments, the CODE 30 material was 
intermixed with the previously utilized common toxic 
lithotype in application schedules ranging from 0.50 
wt% to 5.00 wt% at 0.25 wt% intervals. 

The entire bench scale experimental effort 
consisted of a total of 3285 individual experiments 
with 65,700 analyses. This point is made only to 
emphasize the fact that the database is sufficient- 
ly large to allow a reasonably high statistical 
reliability to be assigned to any effectivity com- 
parisons that will be made. 

SMALL FIELD SCALE (BARREL) EXPERIMENTS 

The second phase of this work involved small 
field scale (BARREL) experiments. The primary 
objectives of this study were twofold: 1) to de- 
termine the effect of phosphate mine waste on acid 
production from bituminous mine waste rock under 
natural conditions of weathering and 2) to provide 
the data needed to calculate conversion factors to 
allow bench scale data to be related to real weath- 
ering situations. The fact that phosphate rock can 
significantly reduce the production of AMD had al- 
ready been established by the authors by previous 
experimentation. However, statistical analysis of 
the data showed that the effectiveness of phosphate 
was not directly related to any stoichiometric for- 
mulation based on the percentage of pyrite in the 
rock sample. Therefore, the ameliorative property 
of phosphate must be mass transfer controlled. 
This being the case, the sample treated with the 
smaller particle size, but equal application rate 
of rock phosphate will be more effective as an 
ameliorant . 

The soxhlet experiments tested the effective- 
ness of surface area and weight percent of phos- 
phate on 100 gram samples of toxic material. Be- 
cause of the simplicity of this experimental pro- 
cedure, large numbers of experiments can be run and 
a large data base can be accumulated. The question 
that yet remained to be answered was how these 
soxhlet tests related to actual atmospheric weath- 
ering. 

The toxic materials used in these experiments 
were of two types: 1) a cleaning plant waste from 
Grant County, West Virginia, the toxic "standard" 
material and 2) a sandstone overburden rock of the 
Kittanning coal bed in central West Virginia. Com- 
pared to many toxic rock wastes studied over the 
past four years, the cleaning plant waste material 
is of average toxicity. The Kittanning coal beds 



are some of the most extensive in the state. As a 
result, the mining community must frequently con- 
tend with the sandstone overburden material. Six- 
teen tons of each material were obtained and 
screened to a top size of 1.5" in order to guaran- 
tee that when packed into the plastic containers, 
the water would not channel through the barrels but 
rather would travel downward in a uniform front. 

Identical sets of the samples and correspond- 
ing treatments used in the field scale experiments 
were prepared and evaluated by the soxhlet bench 
scale procedure experiments to provide the data by 
which the field scale and bench scale data could be 
correlated. 

Based upon the finding from the bench scale 
experiments that rock phosphate particle sizes 
greater than 1/16" showed limited ameliorative 
effectiveness, only to fine sizes were used in the 
barrel experiments. For the small field scale ex- 
periments, the Code 30 and 31 materials were fur- 
ther sub-sized into four seive size ranges: 1) 
18-35 mesh, (750 microns), 2) 36-60 mesh, (375 
microns, 3) 60-120 mesh (187 microns), and 4) 
less than 120 mesh (125 microns). 

A total of ninety field scale barrel experi- 
ments were prepared (see Table 4) . Each experi- 

TABLE K S.1ALL FIELD SCALE {BARREL) EXPER IJ!iL'J7S 



MATERIAL 


AIIELI0RAI1T 


PARTICLE KSSH SIC 


APPLICATION RATL RE 

1.20S 1 
1.205 2 
1. 205 3 


; ::uhber 


RECUSE 

REFLSE 
REFUSE 


P-ROCK 

P-ROCK 
P-H0CK 


IP -35 

18-35 
18-35 


01 
02 
03 


REFUSE 
PEFJSE 
REFUSE 


p-rock 

P-RCCK 
P-ROCK 


lP-35 
lfi-35 
18-35 


0.005 1 

c.4o:. 2 
o.uoi 3 


04 
05 
06 


REFUSE 
REFUSE 
REFUSE 


P-ROCK 
P-ROCK 
P-ROCK 


IP -35 
18-35 
1P-35 


0.133 1 
0.135 2 
0.135 3 


07 
08 
09 


REFUSE 
KEFL'SE 
REFUSE 


P-ROCK 
P-ROCK 
P-ROCK 


13-35 
18-35 
18-35 


0.045 1 
O.aaS 2 
O.oes 3 


11 
12 


REF'JSE 
REFUSE 
HEFUSE 


P-RUCK 
P-ROCK 

P-ROCK 


35-60 

35-60 
35-60 


1.205 1 
1.205 2 
1.205 3 


13 

10 

15 


REFUSE 
REFUSE 
REFUSE 


P-ROCK 
P-ROCK 

P-ROCK 


35-60 

35-60 
35-60 


0.405 1 
O.uCS 2 
0.405 3 


16 
17 
IB 


RE.-'JSE 
REF'JSE 
REFUSE 


P-ROCK 

P-ROCK 
P-ROCK 


35-60 

35-60 
35-60 


0.135 1 
0.135 S 
0.135 3 


19 

20 
21 


REFUSE 
REFUSE 
REFUSE 


P-K0CK 
P-ROCK 
P-ROCK 


35-60 
35-60 
35-60 


0.005 

0.015 S 

0.0*5 3 


22 
23 

2o 


REFUSE 
REFUSE 
RSPUSE 


P-ROCK 
F-K0CK 

P-KOCK 


60-: 20 

60-120 
60-120 


1.205 1 
1.205 S 

1.205 3 


25 
26 
27 


REFUSE 
REFUSE 
REFUSE 


P-ROCK 

P-RCCK 
P-ROCK 


60-120 
60-120 

60-120 


o"«05 

0.105 3 


28 
29 

30 


REF'JSE 
REFUSE 
REFUSE 


P-RUCK 
P-ROCK 

P-ROCK 


60-120 
50-120 

60-120 


0.135 1 
0.135 i 
0.135 3 


31 
32 
33 


REFUSE 
REFUSE 
REFUSE 


P-ROCK 

P-ROCK 
P-ROCK 


60-120 
60-120 
60-120 


0.045 
0.005 

0.005 : 


34 
35 
36 


REF'JSE 
REFUSE 
REFUSE 


P-ROCK 
P-ROCK 
P-ROCK 


LT 120 

LT 120 
LT 120 


1.205 

1.205 
1.205 


1 37 

2 3e 

i 39 


REFUSE 
REFUSE 
REFUSE 


P-ROCK 

p-rock 

P-ROCK 


LT 120 
LT 120 
LT 120 


0.405 
0.005 
0.005 


1 40 

2 *1 

3 42 


REFUSE 
HEFUSE 
REFUSE 


P-ROCK 

P-ROCK 
P-ROCK 


LT 120 
LT 120 
LT 120 


0.135 

0. 135 
0.135 


1 «3 

3 45 


REF'JSE 
REFUSE 

REFUSE 


p-nocK 

P-ROCK 
P-HOCK 


LT 120 
LT 120 
LT 120 


0.005 
0.0*5 
0.005 


1 06 

2 «7 

3 48 


REFUSE 
REFUSE 

RErUSE 


P-SLURRY 
P-SLURRY 
P-SLURRY 





0.805 
4. 80S 
0.605 


1 49 

2 50 

3 51 


REFUSE 
REFUSE 

REFUSE 


P- SLURS* 
P-SLURRY 
P-SLURRX 




1.605 
1.605 
1.605 


1 52 

2 53 

3 5" 


REFUSE 
REFUSE 
REPOSE 


P-SLfRRt 
P-5-URSX 
P-SLJKRY 




0.525 

0.525 
0.525 


1 55 

2 56 

3 57 


REFUSE 

REFUSE 
REFUSE 


P-SLURRY 
P-SL-'HRY 

P-SLUKRY 




0.165 
0.165 
0.165 


1 56 

2 59 

3 60 


REFUSE 
REFUSE 
REFUSE 


COIITR0L 
CQH7S0L 
COIlTnOL 






1 61 

2 62 

3 63 


SAIJDJTO 
SAI.DSTO 
SAIIDSTO 


IE P-ROCK 
iE P-ROCK 
X P-ROCK 


16-35 
13-35 
1P-35 


0.405 
0.1*05 
0.405 


1 70 

2 71 

3 72 


3aud570ke p-rock 
sa;;dsto;je p-rock 
saj;dsto::e p-rock 


35-60 
35-60 
35-60 


0.1.05 
0.405 
0.405 


1 73 
3 75 


SAIIDSTO 
SAIQSTO 
SAIIDSTO 


IE P-ROCK 
JE P-ROCK 
IE P-ROCK 


35-60 
15-60 
35-60 


0.005 
0.005 

O.ool 


1 76 

2 77 

3 78 


SANDCTOIIE P-H0CK 
Sft!.DS70:.E P-HOCK 
SAKD5T0KS P-ROCK 


LT 120 

LT 120 
LT 120 


0.405 

0.405 
0.405 


1 79 

2 90 

3 81 



70 



SAKHSTOIIE 
SATOSTOME 
SAIJOSTOJJE 


P-SLURRY 

P-SLUfUlY 

P-SLURRY 


••80S 1 
«.S0S 2 

-.so-. 3 


SAHasrcrir. 


P-SLURRY 

P-SLURRY 

P-SLURRY 


1.602 1 
J. 605 2 
1.60X 3 


sakdstone 

sacoetohe 
sakdstoi.t 


P-SLURRY 

P-SLURRY 


0.52S 1 
0.525 2 
0.525 3 


SAiosrom 

5AKSS70NE 

SAJ3S70HE 


P-SLURRY 

P-SLURRY 

P-3LURRY 


0.165 1 
C.16S 2 
0.161 3 


SANTS70J.E 
SANDSTONE 
SAHDSTOIIE 


COHTROL 

CONTROL 

COIfTROL 





ment consisting of three hundred pounds of toxic 
material with the appropriate phosphate ameliorant. 
The barrels were placed on 8 benches, with 12 
experiments per bench; the placement of individual 
experiments was determined by random number gener- 
ation. 

TREATMENT OF THE DATA 

The basic data collected from the leachates of 
both the bench scale and small field scale experi- 
ments included pH before and after the addition of 
hydrogen perioxide, specific conductivity, acidity, 
the concentrations of Ca, Mg, Na, K, Fe, Mn, Al, 
Si and sulfate. The sulfate concentration was used 
as a measure of original acid production and was 
tabulated in cumulative milligrams of sulfate 
after each cycle of the bench scale experiments. 
From these data and the initial sulfur content of 
the rock material, a percent unreacted sulfur (%Su) 
remaining within the rock material after each 
leachate collection was calculated. 

Because laboratory leaching and field leaching 
were conducted under totally different experimental 
conditions, it was necessary to equate a 14 day 
soxhlet cycle to the actual field days of the bar- 
rel experiments. To accomplish this, the number of 
field days required to reduce the sulfur content of 
the rock material to the same percent unreacted 
sulfur (%Su) for each of the four cycles of the 
bench scale data were calculated. These data were 
plotted versus the corresponding times for the 
bench scale data (figure 4) . A regression relation- 



scale data to be equated relative to real time. 

EXPERIMENTAL PARAMETERS 

Previous research by the authors had estab- 
lished four experimental parameters which were 
used to evaluate the toxic potential of an acid 
producing system: 1) an acid production rate con- 
stant, ALPHA, which is the slope of the plot of the 
naturallogarithm of the percent of unreacted sul- 
fur of the toxic material versus real time (see 
fig. 5) 2) a sulfate production parameter, S300, 




20 30 40 60 

Tlmo (In Lab Days) 



FIGURE S. Percont Unreacted Sulpnur C%Su) Ver 
Days lor Banc!) Scalo Experiments 



which is the calculated amount of sulfate produced 
from 1000 Tons of toxic material in 300 days. This 
parameter is used as a comparison statistic, 3) 
the ACID Load which is the equivalent amount of 
CaC03 Tons of the toxic material and 4) a parameter 
ETA, which is calculated by dividing the acid load 
generated from 1000 Tons of the toxic material by 
the total sulfur content of the toxic material. 




10 20 30 40 60 

Laboratory Day* (Bench Seal* Data) 



FIGURE 4. Time to Reduce % S to Eouel Concentratlona of 
Unreected Sulphur 



ship was used to calculate the real time days re- 
quired to achieve the corresponding %Su observed in 
the laboratory experiments. The results showed that 
one 14-day laboratory cycle was equal to 8.8 real 
time days. The ratio 8.8/14 was therefore the 
scaling factor which allowed all subsequent bench 



RESULTS OF THE INVESTIGATION 

Bench Scale Experiments 

The results of the investigation are quite 
clear. The effectiveness of the rock phosphate is 
simply a question of mass transfer; the effective- 
ness increasing with decreasing particle size and 
increasing magnitude of the application rate. 

The results of experimental SET #1 are illus- 
trated in Figure 6. The toxicity parameters for 
the rock material used in the individual experi- 
ments is indicated on the figure for each particle 
size experiment. It is significant to note that 
the sand to pebble-sized material is relative in- 
effective at acid reduction. Even the -1/16" size 
does not reduce the acid load by 50% at projected 
application schedules of 5 wt%. 

The CODE 30 material, as would be expected 
was significantly more effective with 8 out of 12 
experiments resulting in a 50% acid reduction with 
an average application schedule of about 2.2 wt% 
apatite. Of the remaining 4 experiments which 
showed relatively low response, the sulfur content 
of the toxic material in two of the experiments 
was in excess of 9 wt%, thereby explaining the 
limited response. 



71 





IQO- 
















CODE 30 


fltlS 




80- 


CODE 


30 


V 


... 


■ / 






*^CO0E 30 

1 


- ■ -0.0004 


o 


eo - 
















y 




K 






















< 
* 


40- 
20 - 
0- 










1/16* - 
1/4* 


It 


:- 


« 1/16* 

i/a* - i/4* 


tuE 























Wt. % Apatite Addition 

FIGURE 6. Percent Acid Reduction va. Wt. % Apatite Application 
for Various Sized Rock Phosphata Materials 

Most encouraging, however, were the results of 
experiment SETS #2 and #3 which utilized the dried 
slurry material. Because the material contained 
only about 25% by weight apatite, the effectiveness 
of the material used without any apatite addition 
was limited (fig. 7). However, with the addition 



PHOSPHATE STUDY 

To«lc Material LO. - CC6434 

Aerlco Dry P0 4 
P0 4 Material LD. - cc 11 ee 



Wt. % Apatite Addition 



FIGURE 7. Percent Acid Reduction va. Wt. % Apatite Application 
for Dried Slurry - No Code 31 Spike 



of CODE 31 material, the slurry material showed 
significant improvement in performance. The re- 
sults of experimental SET #2 are summarized in 
figures 8 and 9 which plot acid load and alpha re- 
duction versus wt% apatite addition respectively. 



1.1- 


\ 

• 




1.2- 


\* 






• \* 


• 


1.1- 


\ 


• 




•\ 




1.0- 


\ 


• 


O.a- 




\ .. 
\ 




• 


0.8' 






0.7- 




\ 
\ 


o.«- 




V 


0.5- 




V 


0.4' 




\ 


0.3- 




_ aNa 


0.2- 




• .N 

>» e 


0.1- 




• * ~~ — — 


0.0- 








1 


1 I I 1 1 

2 3 4 s e 



App. Rata tn Wt. % Apatlta 



FIGURE a. Po-cont Acid Reduction va. Wt. % Apatlta Application 
for Dried Slurry with Coda 31 Splka 



1.5- 


\ 

a 






1.4- 


\. 
\ 






_ 1.1' 

o 


\ . 
• \ 




e 


1 1.2- 
U 


\ 






o 1.1- 


. \ 


\ 








\ 




S 1.0- 




a 


< 




•\ 










- 0.9- 






\ . 


• 






\ " 


i °- 8- 






\ 


eg 






. \ . * 


? 0.7- 






\ • 


■ 






\ 


• O.o- 






\ 


° 0.5- 






\e 








1 0.4- 






*• 


l 0.3- 














-V 


= 0.2- 






1 V . . 


0.1- 






. •--•.-_ 


0.0- 










> 


i 


2 3 4 5 7 



App. Rate In wt. % Apatite 

FIGURE 9. Percent Alpha Reduction va. Wt. % Apatite Application 
for Dried Slurry with Code 31 Spike 



The data show that approximately 0.9 wt% apatite 
must be added before acid reduction is initiated. 
Below 0.9 wt% apatite addition, the chemical 
effect of calcium on the system is apparently the 
same as observed for calcite addition, namely, it 
results in an increase in acid production. Be- 
tween 0.9 wt% and 4.0 wt% apatite addition, how- 
ever, reduction in acid production is rapid. Note 
that on the average, an application schedule of 
5 wt% apatite reduced both acid load and the rate 
of acid production by more than 90% relative to 
the untreated control. As previously shown, cal- 
cite must be added at a 5 wt% addition schedule 
simply to INITIATE positive acid reduction. 

The results from the slurry-CODE 31 mixtures 
of experiment SET #3 are illustrated in figures 10 
through 17 and the data from SET #4, the CODE 30 
addition to the "standard" prep refuse, are shown 
in figures 18 and 19. The trends in these data are 
similar to those of the previous experiments. 



72 



FIGURE 10 

ROCK TYPE - ROOF SHA LE 

STRATIGRAPHY - UPPER FREEPORT 

TOTAL GULFUR - 2.410 

ALPHA - 0,0088 

ETA - 25.7 



FIGURE 14 

ROCK TYPE - PREP REFUSE 

STRATIGRAPHY - UPPER FREEPORT 

TOTAL SULFUR - 3.313 

ALPHA - 0.000 



i RATIO 



BLURRY: C0OE3I RATIO 




APPLICATION 



APPLICATION SCHEDULE (pirtl/1001 



FIGURE 1 1 

ROCK TYPE - ROOF SHALe 

STOAT ICRAPHY - UPPER FREEPORT 

TOTAL SULFUR - 2.418 

alpha - o.ooea 



FIGURE 15 

ROCK TYPE - PREP REFUSE 

STRAII6RAPHY - UPPER FREEPORT 

TOTAL SULFUR - 3.312 



SLURRY: COOE31 RATIO 1:0.0 




ioo 4> 

o « 






' bo 








c ~ 








{ - 

i - 




♦^ 




: - 






o " — ■ 


"' 








' 30- 
















N 

T « 








R 

« 








l , . . - ■ 


' J 


' 



APPLICATION SCHEDULE (part*/ 100) 



APPLICATION SCHEDULE (parta/100) 



FIGURE 12 

ROCK TYPE - PREP REFUSE 

BTRATISHAPHY - UPPER FREEPORT 

TOTAL SULFUR - 8.82S 



FIGURE 16 

ROCK TYPE - BHALE PARTING 

STRATIGRAPHY - LOWER FREEPORT 

TOTAL SULFUR - 1.840 

ALPHA - 0.0033 



SLURRY: C0OE31 RATIO 1: O.B 



150- 








1«- 


O >v 






130- 








120- 




Ss x. • 




110- 








100- 


> 






90- 








00- 








70- 








*0 






o ^""^■-^^ 


BO 








40 


















APPLICATION SCHEDULE Inn 



APPLICATION SCHEDULE (parta/100) 



FIGURE 13 

ROCK TYPE - PREP REFUSE 

STRATIGRAPHY - UPPER FREEPORT 

TOTAL SULFUR - 8. 823 

ALPHA - 0.0O9 



SLURRY. COOES I RATIO 1: O.B 



C WO 

I 

110 






FIGURE 17 

ROCK TYPE - SHALE PARTIHO 

STRATIGRAPHY - LONER FREEPORT 

TOTAL SULFUR - 1.040 

ALPHA - 0.0033 




APPLICATION BCHEOULf lpwt«/100l 



APPLICATION SCHEDULE lpWta/1001 



73 



FIGURE 18 



ioi«i- sul/up. - .»■•■ 

ALPHA - 0.0O» 



FIGURE 21 



percent acid reduction 




3. a? 

8.33 

APPLICATION RATE fIOMS/100 TONS) 



APPLICATION SOCIU1 (P«rt( 



FIGURE 19 

ROCK type - pp*p r efuse 

■TIUTIBfUPHV - UPPO* fHEEFORT 

TOTAL SULFUR - *-»» 

ALPHA - 0.007 

ETA * 15.164 

PHOSPHATE 10 - CODE 30 



FIGURE 21 -A 



PBKfNT ACID MDUOTION 




1.33 

0.87 
APPLICATION RATE (TONS/100 TONS) 



APPLICATION SCHEDULE Ip«rt«/t00> 

SMALL FIELD SCALE EXPERIMENTS 

The results of the small field scale experi- 
ments are shown in figures 20 through 21A. Similar 



FIGURE 20 



size CODE 

I ■ 700 ■Icrona 
' - 373 B Micron* 
I • 167.9 ■leron* 
4 - 18S ■Icrona 



APPLICATION RATE (TOMS/100 TOMS) 



VAP.10UB PHOttPHATE PAPT1CLC SIZE 
III ECIO ON ACID PRODUCTION 
OP A S1MSLC TOXIC MATERIAL. 



to the bench scale experiments, the data show the 
ameliorative effectiveness to increase with de- 
creased particle size. The three dimensional plot 
in figure 21 shows however, that when dealing with 
fine particle sizes, the application schedule is 
more influential on ameliorative effectiveness 
than is particle size. Statistically, the amelio- 
rative reaction of all the fine sizes tested was 
similar. Most significant is that both the bench 
scale and field scale sets of experiments show that 
maximum ameliorative effectiveness of rock phos- 
phate is achieved at about 4 wt% apatite addition. 

CONCLUSIONS 

1) Rock phosphate (apatite) is an effective 
AMD ameliorant when used in particle sizes 
less than 18 mesh (750 microns) and with 
application rates in excess of about lwt%. 

2) Maximum ameliorative effectiveness will be 
achieved using fine ground rock phosphate 
at an addition rate of about 4 or 5 wt%. 

3) Below a particle size of 18 mesh, the par- 
ticle size of rock phosphate becomes less 
important than the application rate in 
acid reduction. 

4) The clay slurry spiked with -375 mesh apa- 
tite is an extremely effective AMD 
ameliorant. 

5) Bench scale experiments can effectively 
evaluate and predict the response of toxic 
materials or combinations of toxic mater- 
ials and ameliorants exposed to atmosphere 
weathering. 



74 



ACKNOWLEDGEMENTS 

The authors would like to acknowledge and ex- 
press their appreciation to the W.Va. Dept. of 
Energy and to the Florida Institute of Phosphate 
Research who funded the research reported in this 
paper. We would also like to thank Ralph Chamness 
of Texasgulf Sulfur, Aurora, North Carolina and Hal 
Miller of the Agrico Mining Company of Mulberry, 
Florida who provided the phosphate materials used 
in this study. 

REFERENCES 

Stumm, W. and Morgan, J.J., 1970, Aquatic Chemistry, 
Wiley Interscience, John Wiley and Sons, N.Y., 
p. 522. 

Stiller, A.H., Renton, J.J., and Rymer, T.E., 1984, 
A Mathematical Hypothesis for Determination of 
Mine Effluent Behavior, Proc. Surface Mining 
and Water Quality, 5th Annual West Virginia 
Surface Mine Drainage Symposium, Morgantown, 
WV, March 1984. 

Renton, J.J., Stiller, A.H., and Rymer, T.E., 1985, 
Relationships of Various Lithologies and 
Geologic Strata for Acid Producing Potential, 
Proc. Surface Mining and Water Quality, 6th 
Annual West Virginia Surface Mine Drainage 
Symposium, Morgantown, WV, April 1985. 



75 



A COMPUTER SIMULATION PROBABILITY MODEL FOR GEOCHEMICAL PARAMETERS 
ASSOCIATED WITH COAL MINING OPERATIONS 

Thomas E. Rymer II, John J. Renton, and Alfred H. Stiller^ 



Abstract. — There have been many attempts to empirically 
model the environmental effects of various geochemical parame- 
ters associated with mining. Each of these attempts has en- 
countered the single greatest barrier of any empirical approach; 
the randomness of data. Randomness manifests itself in many 
forms from the degree of variability within a specific dataset 
to the degree of intrinsic error associated with the measure- 
ment of certain variables. There are simply too many random 
processes, variables, and interrelations associated with a coal 
mine site to allow the depiction of any environmental response 
within reasonable certainty using only basic scientific princi- 
ples, equations, and empirical formulae. Computer simulation 
modeling provides an effective mode of evaluating the intrin- 
sics of such a random system. In order to generate a random 
probability simulation model, it is necessary to isolate perti- 
nent elements of the system, which in this case are sulfate con- 
centration, effluent flow, and kinetic rock properties. Some 
logistical set of formulated interactive governing rules must 
be then devised whereby the model can be limited to those as- 
pects of the system which are deemed to be pertinent to the 
analysis and types of solutions for which answers are sought, 
which in this case is simultaneous pyrite oxidation-sulf ate 
elimination. Within the context of this logic, a quantifiable 
ratio between the rate of acid generation within a mining 
system and the rate of acid elimination from the system has 
been derived. With this information in hand it is now possible 
to project the longevity of acid effluent from a mining opera- 
tion, develop better treatment or amelioration strategies, and 
determine the chemical impact of these mining operations on the 



immediate localized watershed. 



1 Paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and Rec- 
lamation and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

2 Thomas E. Rymer II is a Research Associate, Dept. 
of Geology and Dept. of Chemical Engineering, West 
Virginia University, John J. Renton is a Profess- 
or, Dept. of Geology, West Virginia University 
and Geochemist, West Virginia Geological and Eco- 
nomic Survey, and Alfred H. Stiller is an Assoc- 
iate Professor, Dept. of Chemical Engineering, 
West Virginia University. 



76 



INTRODUCTION 

A mathematical model is an attempt to quanti- 
fy a specific parameter associated with a system on 
the basis of other measurable variables. The ma- 
jority of such models deal with limited aspects of 
mining such as hydrological predictions, level of 
acid generation, or acid amelioration. Most of the 
models can be classified as being empirical; that 
is, a model dealing with variables simplified to 
standard regression formulae which usually have a 
basis in some standard chemical or engineering prin- 
ciple. A macromodel on the other hand, is a model 
which draws conclusions or predictions based on the 
entire mine operation with all of its interrela- 
tions and interactions. 

Two important criteria exist concerning error 
and randomness and how these factors can effect the 
success of an empirical model: 

1) THE MODEL MUST BE ABLE TO ABSORB THE INTRINSIC 
AND INHERENT ERROR OF THE SYSTEM 

The model cannot be so filled with variables, 
constants, and complexities that the sum total of 
intrinsic error encountered in the measurement of 
the individual component variables and the error 
inherent in any subsequently predicted parameters 
becomes disguised in the form of high uncertainty. 
The inclusion into the model of any single variable 
or group of variables for which there exists a po- 
tentially high intrinsic error of measurement for 
the sake of higher statistical significance 
cannot be justified. 

2) IT MUST BE ABLE TO MINIMIZE RANDOMNESS 

The model may be highly dependent on variables 
which are random by their own nature. This random- 
ness must be absorbed into the model and thus be 
reduced rather than subtly concealed in a region of 
high uncertainty brought about by the randomness. 

A large-scale macromodel for mining operations 
seems to have eluded empirical modeling attempts. 
One need look no further than the two model quali- 
fiers previously listed to discover the reasons; 
namely randomness and variable error. 

Randomness and Mining Operations 

Sources of randomness in an acid producing 
mining system can be found everywhere. Some of the 
sources show high levels of randomness while others 
will show low levels of randomness. A few of the 
highly variable parameters are summarized below: 

1) THE VARIABILITY FOUND IN THE PERCENT TOTAL SUL- 
FUR OF INDIVIDUAL OVERBURDEN LITHOTYPES 

2) EFFLUENT FLOW MEASUREMENT 

3) CHEMICAL CONCENTRATION VS EFFLUENT FLOW The 

plot shown in figure 1 illustrates a mathe- 
matically random relationship between efflu- 
ent sulfate concentration and flow 

4) CLIMATIC FACTORS 

5) TURBIDIMETRIC SULFATE ANALYSIS used as the 

measure of a acid generated 

6) VARIABILITY OF PYRITE AS DETERMINED BY X-RAY 
DIFFRACTION ANALYSIS 

7) VARIABILITY OF THE KINETIC RATE CONSTANT FOR 
PYRITE OXIDATION 

8) VARIABILITY IN THE ACTUAL CHEMICAL MODEL 



* i * * 
9 •• -:: #, «. 


i 


♦ 










•• • « i • • 

♦ 


i 


! • 

• 
♦ 


1 


♦ 


* 


• 


■ MM 




M 








M 


rut 


(i»l/«ln| 










Mgur* 1. flandoa nature of luldt 
Data Itom alt* dui 139- 


t coneanttatl 
Ft. 




fie*. 







DESCRIPTION OF RESEARCH 

A Random Probability Simulation Model 

An alternative solution to the problems synono- 
mous with empirical modeling is to develop a logical, 
quantifiable model that absorbs both randomness and 
error. Such a modeling scheme exists in Probability 
Simulation Modeling (PSM) . In this type of predic- 
tive modeling, the randomness of the data itself 
and the error associated with individual measure- 
ments are both absorbed into the probability dis- 
tribution of the data. The difference between the 
PSM approach and empirical modeling is that in PSM 
modeling the DISTRIBUTION of the data rather than 
the magnitude of INDIVIDUAL data points form the 
working data set. 

One needs then to develop a logical set of 
rules to govern the relationship of the data dis- 
tribution and any experimental parameter of inter- 
est. It is therefore possible to evaluate the ef- 
fects of numerous possible combinations of param- 
eter values (data randomness) that exist in their 
respective distributions. 

Two important properties of a PSM are: 

1) the scientific or mathematical principles gov- 
erning the system have been obeyed and, 

2) the statistical integrity of the original data 
is upheld. 

The objective of this work was to develop a 
model which would focus on the two basic processes: 
1) the production of acid within a minesite and 2) 
the elimination of the acid from the system. A 
well-established mathematical formulation known as 
Simultaneous Species Production-Elimination (SSPE) 
exists that logically describes a system. In ess- 
ence, this principle states that the change in the 
rate of the appearance and disappearance of any 
species from a system is given by the differential. 



dS 



dt 



= aS(t) - bS(t) 



(EQUATION 1) 



This mathematical relationship states that the 
rate of change with time of any chemical parameter, 
(S) , is equal to the rate of production of S, (a), 
times the amount of S present at some time (t), 
minus the rate of elimination of S, (b) , times the 
amount of S present at the same time, t. 

This differential equation is then solved to 
yield a working mathematical relation: 



77 



-at 



S(t) 



(b - a) 



( e 



-bt 
e ) 



(EQUATION 2) 



This is a member of a series of differential 
equations called the "Bateman Equations', after the 
man who derived them. 

The rate constants a and b will subsequently be 
referred to as alpha and beta, respectively. 

The SSPE model is utilized quite frequently in 
the study of natural systems such as the determin- 
ation of the abundance of an intermediate radio- 
active isotope in a breakdown sequence, biochemical 
assimilation of hormones, and in the study of the 
potential effects of toxic wastes on biological 
systems . 

The SSPE equation offers an excellent governing 
principle for a systems model that will allow the 
monitoring of pyrite oxidation products (including 
acid effluent) associated with a mining operation. 
The analogies to radioactive decay and biochemical 
production assimilations are parallel. In a mining 
operation, acid is formed as pyrite oxidizes to 
sulfate at a rate alpha. As a result of ground 
water movements, the sulfate is removed from the 
minesite at a rate beta. The overall process can be 
written: 



alpha 



beta 



FeS 



SO (in site) 
4 



SO (effluent) 
4 



Once the values of alpha and beta (production 
and elimination) of the sulfate species are known, 
questions concerning the products of pyrite oxida- 
tion that can be answered include: 

1. How fast are they being formed? 

2. How fast are they being eliminated? 

3. How long will the acid-producing process last? 

4. What will be the effect on the system of a re- 
duction in the magnitude of alpha as a result 
of the incorporation of an ameliorant? 

These are questions that need to be addressed 
in order to solve the problems of acid mine drain- 
age (AMD) , but are questions that cannot be answer- 
ed using empirical models without the introduction 
of large uncertainties. However, through a prob- 
ability simulation analysis, solutions to such 
problems can be found. 

The Generalized Concept of AMD as a SSPE System 



LFATE IN SYSTEM 




TIM lift «•?■) 



sulfur available in a mining system that through 
time exists as either unreacted pyrite, sulfate 
built up in the minesite, or discharged from the 
minesite as effluent. Note that there is a build- 
up of the products of oxidation (sulfate) in the 
system from the outset. The magnitude of alpha 
will determine the rate at which these oxidation 
products are generated. Conversely, beta is the 
prime indicator of the extent to which these pro- 
ducts are being flushed from the system. It is, 
therefore, the ratio of alpha to beta that provides 
a complete insight into the buildup of oxidation 
products in the system. The equation tends to show 
that the time required for pyrite oxidation is 
short compared to the time required to flush the 
oxidation products from the system. The PSM pro- 
vides a means to quantify these observations. 

The Development of a Simulation Model 

The goal of a simulation is to provide a math- 
ematical procedure that will: 

1) be able to utilize analytical data to predict 
parameters of an acid-producing system, 

2) be as unaffected as possible by data randomness, 

3) adhere to sound scientific principles, and 

4) reproduce the statistical distribution observed 
in actual field data. 

In order to develop such a simulation, it is 
necessary to know: 



1) 
2) 



3) 



mine size and geometry, 

field sulfate concentration and flow measure- 
ments of the effluents, the date and the dis- 
charge point where these measurements were 
made, and 

geochemical and chemical kinetic rock proper- 
ties (Sulfur content and Alpha values associ- 
ated with the strata encountered in the mine- 
site) . 



The water quality and flow data used in this 
study were obtained from three sources: 1) Data 
from samples collected at the actual field sites 
and analyzed by the Analytical Section of the West 
Virginia Geological Survey. Effluent flow measure- 
ments were made at the time of sample collection. 
2) Data compiled in the Office of Surface Mining 
and the West Virginia Department of Energy, and 3) 
Information available on CRIS (Coal Reclamation 
Information System) DATABASE. The CRIS DATABASE 
includes a suite of water quality, overburden 
analyses, and general mine information for more than 
300 minesites in the 12 northern coal producing 
counties of West Virginia. In addition, a complete 
suite of lithotypes associated with coal represent- 
ing five different coal beds over a five-county 
area were collected and analyzed for numerous 
chemical properties, among which were the kinetic 
rate constant, alpha using a method developed and 
reported by the authors, and the total percent 
sulfur. These data are contained in a formatted 
form identified as the "WRI DATABASE OF CHEMICAL 
CHARACTERISTICS OF TOXIC ROCK MATERIALS". A com- 
pilation of the alpha value and the total percent 
sulfur ranges for different lithologies is shown 
in Tables 1 and 2 respectively. 



Figure 2 shows the percentage of original pyritic 



78 



BONE COAL 
OVERBURDEN SHALE 
fIT BEFUSE 
PREP REPOSE 
ROOF SHALE 
SANDSTONE 
SEATEARTH 
SHALE PARTI.'IO 
SILTSTONE 



HIOH 

BANOE 

ALPHA 

0.0030 26 
0.009978 
0.O1O6B1 
0.030556 
0.026199 
0. 13 in 20 
0.167*21 
0.105791 
0.02U8A 



LOU 
RAIICE 
ALPHA 

0.00027* 
0.000073 
0.001020 
0.00107' 
0.000165 
0.000271 
0. 001*28 
0.000676 
0.006730 




PUH Kal/Btnl 



TYPE 


LOU 


HIOH 




BANCE 


RANOE 




SULFUR 


SULFUR 


BONE COAL 


0.195 


5.000 


OVERBURDEN SHALE 


0.221 


9.736 


PIT REPOSE 


0.080 


• .202 


PREP REFUSE 


0.032 


9.572 


ROOF SHALE 


0.030 


7.010 


SANDSTONE 


0.001 


10.063 


SEATEARTH 


0.037 


9.1«« 


SHALE PARTIHO 


0.0»7 


6.33« 


SILTSTOHI 


0.0*6 


0.37« 



A coaputsr aanaratad statistically significant 
ralatlonahlp rot aultats concentration va flo- 



The Simulation 

Following is a brief description of an actual 
computer simulation defining what the computer is 
doing and what decisions the computer is expected 
to make. 

1. The computer reads into memory the following 
minesite data: 

The permit number 
Dates of active mining 
Present mine status 
Type of surface mine 
The disturbance acreage 
The computer reads into memory the following 
field data: 

date of sampling 
location of sampling 

sulfate concentration of the effluent 
effluent flow measurement 
3. The computer then determines the statistical 
relationship between the sulfate concentra- 
tion data and the flow data for each samp- 
ling location. This is done for each 
season of the year. It is at this point 
that the computer is programmed to decide 
whether the sulfate-flow relationship for 
each discharge point and each annual season 
is: 
a. A completely random number (Figure 8) 



The 
The 
The 
The 



I ' • 



* * • 



FLAW Hal/»KO 



■castration va (law. 






'■H«»[i I. H.ndcm niturt of iu 

Dili ftoa «Ua BLM ■- 

A statistically significant inverse 
curve (Figure 9) 



untoied during tho study. 

c. A multimodal random number (Figure 
10) 




These are the three relationships that have been 
shown to exist between sulfate concentration in 
the effluent and the flow measurement of the 
effluent. The criteria on which the computer 
makes this decision can be seen in table 3. 

Table 3. Determination of data distribution type based on flow and 
sulfate concentration data. 



TYPE OP DATA 
DISTRIBUTION 


TYPE OP 
CORRELATION 


CORRELATION 
COEFFICIENT 


REMARKS 


TOTALLY RANDOM 


PEARSON 

H0I-60J OP 
10 RANDOM POINTS 
SHOW NE0ATIVE SLOPES 
IN 1000 TRIALS 


< 0.20 


(SEE PI0. 8) 


INVERSE 


POKER 
TRANSFORMATION 


> 0.90 


(SEE PI0. 9) 


MULTIMODAL 

RANDOM 


CANONICAL 
QROUP CLUSTER 




DEFAULTS IP 

OTHER 

DISTRIBUTIONS 

FALL 

(SEE PI0. 10) 



A. The computer will then read available strat- 
igraphic information, determine the volume 
percent of each significant stratigraphic 
layer, and will then determine an operational 
alpha value for the minesite based upon the 
compilation of alpha values in the "WRI 
DATABASE" and the geologic cross section or 
drill core log. 

When all of the aforementioned information has 
been generated, the computer enters "RANDOM SIMULA- 
TION MODE", which is the portion of the software 
that determines the BETA value associated with the 
minesite . 



79 



RESULTS AND CONCLUSIONS 

Results of the Simulation 

Figure 3 is a histogram of beta values gener- 
ated by 1 million computer simulations and beta 



actual (field) and 




generations for a particular minesite. The field 
data for this graph are from the DLM permit 135-7J 
in Upshur County, WV. This distribution of beta 
is quite acceptable based on the narrow range of 
potential beta values that were generated. 



0.0001 0007 0.0000 
MTA VALUI 



o.th.) Road 



Figure 4 shows a histogram for 1 million com- 
puter-generated beta values that correspond to field 
data on the Bethel Road gobpile, Monongalia, Co., 
WV. Note that the range of beta values calculated 
for the gobpile is smaller than the range of beta 
values for the DLM 135-78 site. The reason for the 
difference lies mainly in the characteristics of the 
materials of which the two sites are constituted. 
In a gobpile the material is essentially a single, 
relatively homogeneous lithology, namely prep plant 
refuse, while in a reclaimed surface mine, there is 
a diversity of lithologies. As a result, the num- 
ber of combinations of possible sulfur and alpha 
values associated with the diverse lithologies 
found in a reclaimed surface mine are considerably 
greater than in the monolithologic gobpile. 

This successful simulation was accomplished 
with a minimum number of empirical formulae. The 
simulation has completely absorbed all of the ran- 
domness associated with the measurement and distri- 
bution of these data. As can be seen in Table 4, 



DISCHARGE FT 



DISCHABOE PT 



2.589.T 
2. 587. 8 



DISCHABOE PT - 2A 



2.2M.1 

2.238.8 



DISCHABOE PT ■ 2B 



2.633.1 
2.599.9 



1 Sulfate measurements in mg/L end flow measurements In gal/oln. 

the statistical distribution of the sulfate con- 
centration and flow values generated in the random 
simulation matches the statistical distribution of 
the original data set very closely. In other words, 
the mining system has been found to adhere to the 
Simultaneous Species Production-Elimination princi- 
ple. 

With the computer generation of a beta value, 
all the data needed to analyze the pyrite oxidation 
and oxidation product elimination associated with 
the system are available. Using EQUATION 2, 
sulfate concentrations of effluents were calculated 
using a time-series algorithm. 

Figure 5 shows the computer-projected sulfate 



70 00 00 



iw its leo 



figure i. Curve enovlno th# lonq.vlty (or alt* OLA 135-76. 

concentration of the minesite discharge of DLM per- 
mit 135-78. The discharge is projected to 250 
mg/L sulfate concentration (drinking water standard). 
From this plot it is possible to make an estimate 
of the acid-producing lifetime of the minesite. 
The computer-generated curve is based on all of the 
sulfate and effluent discharging from a single 
point. Figure 6 displays a significant break that 



80 




UPPERMOST BOUNDARY OF SECULAR EQUILIBRIUM 




10 M M 40 



ot a Blalng (y«t«*. 



occurs when alpha no longer plays a significant 
role in the system. It is at this point that the 
amount of reactive material has been reduced to a 
negligible level. This break point is arrived at 
using the laws of chemical kinetics and the mean- 
lifetime concept of reactive substances. This 
estimation is preliminary and further study is 
necessary to ascertain a more exact break point. 
Table 5 gives the beta and longevity information, 



NINE SITE 


BETA 
1/days 


ACTIVE LONGEVITY 
(years) 


BETHEL R0»D GOBPILE 


> X 


10" 


30-40 


DLM 135-75 NT. TOP REMOVAL 


X 


10" 


' «5 ♦ 


OLN SJ-iS NT. TOP REMOVAL 


' I 


10" 


' «5 ♦ 


DLN 71-75 NT. TOP REMOVAL 


« 


10" 


' «5 ♦ 


DLH 58-77 NT. TOP REMOVAL ■ 


X 


10" 


' »5 ♦ 


EVERETTSVILLE OOBPILE ■ 


X 


10" 


1 25-30 


PIERCE SITE . 


I 


10" 


50 ♦ 


VALHUT HILL OOBPILE l.J 


X 


10" 


28 • 



the time required before the level of oxidation 
products released from the system is environmentally 
acceptable within the context of existing regula- 
tions, for the sites studied. The time (years) axis 
corresponds to the elapsed time from the initiation 
of mining. 

Amelioration and the SSPE/PSM Model 

The amelioration production-elimination model 
allows the development of an amelioration strategy. 
Beta values, being related to physical parameters, 
do not change with amelioration. Alpha, on the 
other hand being related to chemical parameter 
values do change. The purpose of amelioration is to 
reduce the alpha value (the rate of pyrite oxida- 
tion) to a level where the mine system will approach 
a condition known as 'secular equilibrium'. When a 
system has achieved secular equiliblrum, the build- 
up of oxidation products is drastically reduced. 
At this point, the concentration of oxidation pro- 
ducts being generated approaches zero. Figure 7 



shows the theoretical effect on the sulfate concen- 
tration of a mine discharge existing at various 
alpha levels. The alpha and beta values chosen to 
generate this graph are those associated with the 
DLM permit 135-78. The curve is denoted as 
'control'. This graph does NOT show the existing 
effect of amelioration on this particular minesite, 
but what the effect would have been had proper 
amelioration of pyrite oxidation been carried out 
as part of the overall mining operation while the 
minesite was active. The lowest curve represents 
an alpha reduction of 1000 times. The curve 
(alpha = 0.000008) is the upper boundary of 
'secular equilibrium'. The mathematics dictate 
that the sulfate concentration must gradually in- 
crease, but it can be easily seen that below the 
curve corresponding to alpha = 0.000008, in the 
region of 'secular equilibrium', the concentration 
of pyrite oxidation products in the discharge are 
quite manageable. Not only are the concentrations 
of oxidation products being generated very low, but 
the build up of oxidation products in the system is 
also low. 

Hydrological Impact Assessment and the SSPE/PSM 
Model 

The proposed model can also be used to evalu- 
ate the hydrological impact assessment potential 
of a proposed minesite. Knowing the alpha, beta 
and discharge distribution for a particular mining 
operation, a discharge simulation is possible. 
When this discharge simulation is added to existing 
stream quality data and principles of mass balance, 
a hydrological impact simulation (assessment) is 
possible . 

Data were extracted from the CRIS DATABASE. 
Search of the data base located an area where water 
quality and flow data existed prior to the 1982 
initiation of a proposed downstream mining opera- 
tion. This particular mine site was chosen for 
three reasons: 1) the isolated nature of the pro- 
posed site provided an excellent proving ground for 
the simulation, 2) the proposed site was to be a 
mountaintop removal operation; a mine type for 
which the beta (chemical elimination) constants 
for the pyrite-sulf ate system were reasonably well 
established during the course of this current re- 
search endeavor, and 3) stream quality data for a 
downstream area were compiled in 1985. Using the 
pre-1982 water quality data and the available 
general mine information data, a new prototype 
model was developed for the purpose of simulating 
the mine discharge into the stream from a minesite 
exhibiting the chemical patterns reflective of the 
pre-1982 data. The model included the development 
of alpha and beta values that could be expected 



81 



for species other than sulfate ion. The simulation 
was used to project the change in basic water quali- 
ty parameters that would result from the mining 
operation and predict the water quality that would 
be seen by the year 1985. These predicted data 
were then compared to the actual 1985 data taken 
downstream from the 1982 mining operation. Using 
beta values found to be common to mountain top 
removal operations for the respective ions in 
question, this new prototype computer simulation 
package and a system now comprised of the mining 
operation and the receiving stream, the results 
shown by the 1985 water quality study were accur- 
ately projected (See table 6). 

Table 6. CHIA predlctiee capabilities of ?SH modeling technique 
for • leleeted mining operation. 



CHEMICAL 
SPECIES 


1982 

ANALYSIS 

Img/U 


2-YEAR 

PROJECTON 


1985 

ANALYSIS 

(mg/L) 


SULFATE 


9«0 


•light Increase 
to 1,020 ng/1 


1.030 


IRON 


58 


very slight increase; 
eatloate should be 
(statistically) about 
equal 


US 


ACIDITY 


510 


moderate increase (155) 


575 


KANGANESE 


.05 


no statistical 
change 


.01 



CONCLUSIONS 

This research has established several important 
points: 1) the nature of the data associated with a 
mining operation is too randomly distributed and 
too high in intrinsic error of measurement to be 
used in a reliable empirical model. 2) the distri- 
bution of this randomness and error can be incorpor- 
ated into a viable model provided this model is de- 
signed to deal with distributions of data, rather 
than discrete data points or statistical averages 
assumed to be Gaussian distributions. Such model- 
ing technique is called a "RANDOM PROBABILITY 
SIMULATION" or PSM. In order to utilize a PSM one 
must first find a logical, scientifically sound 
principle that governs the system in question. The 
mathematical concept of Simultaneous Species 
Production-Elimination meets all logistical and 
scientific criteria as a governing principle for 
the generation and discharge of geochemical param- 
eters associated with pyrite oxidation in a mining 
operation. 3) this mathematical principle incor- 
porated into a PSM allows the critical evaluation 
of the pyrite oxidation-acid elimination relation- 
ship that exists in a mine. Knowing this relation- 
ship allows scientifically sound amelioration 
planning and with modification, will allow the 
development of a viable hydrological impact 
assessment procedure that will be relatively free 
from guesswork and random error. The relationship 
between pyrite oxidation and effluent discharge for 
a particular mining operation can be quickly 
established using existing field data, making the 
model quite useful in reclamation planning with 
abandoned mine lands. 

ACKNOWLEGEMENTS 

The authors wish to acknowledge the West 
Virginia Department of Energy and the Water Re- 
search Institute, WVU for providing the funding to 
carry on this research. Particular gratitude 
should be given to Mr. Roger Hall who has been a 
constant supporter of our work. 



82 



CHEMICAL STABILITY OF MANGANESE AND OTHER METALS 
IN ACID MINE DRAINAGE SLUDGE 1 



George R. Watzlaf 2 



Abstract. — Federal regulations require mine operators to 
reduce the average concentration of manganese in the effluent to 
2 mg/L. To meet this standard, the majority of mine operators 
add an alkaline material, typically lime or sodium hydroxide, to 
raise the pH to about 10.0. Our laboratory tests using actual 
acid mine drainage (AMD) containing iron, manganese, nickel, 
copper, zinc, and chromium, have indicated that the high-pH 
precipitation method is effective at removing these metals. 
However, manganese, nickel, zinc, and copper in the precipitated 
sludge were susceptible to dissolution upon subsequent 
depression of pH. Up to 30 percent of the original manganese in 
the high-pH precipitated sludge dissolved at pH 7.5. At pH 6.0, 
this figure increased to 78 pet. Additionally, 30 percent of 
nickel, zinc, and copper redissolved at pH values of 6.8, 5.7, 
and *1.7i respectively. Iron and chromium were stable down to 
pH 3.5. No differences were found between the use of lime or 
sodium hydroxide regarding the stability of the precipitated 
sludge. In mine waters containing high concentrations of iron 
in relation to manganese and other metals, manganese and other 
metals were removed at lower pH values than in mine waters with 
less iron. However, manganese in the sludge precipitated from 
the high iron water was less stable. Aging also affected the 
stability of manganese in AMD sludge. After sludge had aged for 
three months, 50 percent less manganese was released upon 
depression of pH. The problem of manganese redissolution can be 
avoided by using an oxidizer such as sodium hypochlorite or 
potassium permanganate; however, these chemicals were 
ineffective at removing copper, nickel, and zinc. To reduce 
manganese below 2 mg/L, the use of sodium hypochlorite or 
potassium permanganate increased chemical treatment costs by 
factors of 2.3 and 2.6, respectively, over the chemical costs 
for sodium hydroxide treatment. 



''Paper presented at the 1988 Mine Drainage and INTRODUCTION 

Surface Mine Reclamation Conference sponsored by 

the American Society for Surface Mining and Chemical Treatment for Manganese Removal 

Reclamation and the U.S. Department of the Interior 

(Bureau of Mines and Office of Surface Mining Federal legislation requires mine operators to 

Reclamation and Enforcement), April 17-22, 1988, adhere to specific effluent limits (U.S. Code of 

Pittsburgh, PA. Federal Regulations 1985 a & b) (table 1). To meet 

2 George R. Watzlaf is an Environmental these limits, treatment of acid mine drainage 

Engineer, U.S. Bureau of Mines, Pittsburgh Research typically involves addition of alkaline material, 

Center, Pittsburgh, PA. commonly lime (CaO or Ca(0H)2) or sodium hydroxide 

(NaOH); natural or mechanical aeration; and 



settling. When the pH of the drainage is raised to 
7 or 8, and there is sufficient time allowed for 
settling, most mine drainage water will meet the 



83 



Table 1 



-Federal effluent limitations 



Pollutant or Maximum Average of daily 
pollutant for any 1 day values for 30 
property consecutive days 



Iron, total * 
Manganese, total 
Total Suspended 
solids 

pH 



7.0 mg/L 
4.0 mg/L 



3-5 mg/L 
2.0 mg/L 



70.0 mg/L 35.0 mg/L 
between 6.0 and 9.0 



*New source performance standards for total 
iron; daily maximum - 6.0 mg/L and 30-day 
average - 3-0 mg/L 



standards for pH, iron, and suspended solids. 
However, this treatment will rarely reduce 
manganese levels below effluent limits because the 
oxidation rate of Mn 2+ is pH-dependent and 
extremely slow below pH 8.5. Diem and Stumm (1984) 
found that dilute solutions of manganese nitrate 
(Mn(N03>2)i maintained at pH 8.4 in the presence of 
dissolved oxygen, showed virtually no oxidation 
for over 4 years. Owens (1963) reported that a pH 
of at least 9.4 was necessary to remove manganese. 
Our tests indicate pH values above 10 are necessary 
for some mine drainage waters. Therefore, to 
ensure adequate removal of manganese, many mine 
operators must add enough alkalinity to raise pH of 
the mine water to about 10. This increases 
chemical treatment costs 20 to 100 pet over the 
costs to remove iron (Watzlaf 1985) and generates a 
large volume of metalliferous Sludge, up to 33 pet 
(by volume) of the treatment plant inflow (Holland 
et al. 1968). Nicholas and Foree (1979) found that 
increasing the treatment pH from 8 to 10 increases 
the required sedimentation basin area from 158 m 2 
to 400 m 2 for NaOH, and from 200 m 2 to 316 m 2 for 
lime. 

The precipitation and removal of manganese in 
an aqueous system is complex. Manganese can exist 
in oxidation states of +2, +3, + 4, +6, and +7 
(Sienko and Plane 1966, Morgan 1967). Initially, 
the valence of manganese in the air-oxidized 
precipitate (removed at high pH) lies between 2.67 
and 3-0 (Mn30ij - hausmannite and B-MnOOH - 
feitknechtite, respectively) (Hem 1981, Murray 
et al. 1985). Both these minerals are eventually 
transformed into the more stable Y-MnOOH - 
manganite (Murray et al. 1985, Stumm and Giovanoli 
1976). Manganese can also be removed by 
adsorption onto reactive surfaces. Ferric 
hydroxide (FetOH^) and manganese precipitates can 
provide surfaces for sorption of Mn 2+ , which occurs 
at lower pH values than necessary for manganese 
removal as hydroxides (Morgan and Stumm 1964, Stumm 
and Morgan 1981, Collins and Buol 1970). There- 
fore, manganese in mine waters with high iron 
concentrations can usually be removed at lower pH 
values than manganese in mine water with lower iron 
levels. However, manganese precipitated with iron 
was found to go back into solution upon 
acidification (Collins and Buol 1970). 



Additionally, manganese can exist in many 
different complexed and chelated forms (Martell and 
Calvin 1952, Clark et al. 1977, Evangelou 1984). 
The formation of manganese complexes depends on 
oxidation state, pH, bicarbonate-carbonate- 
hydroxide equilibria, and the presence of other 
materials (Nalco Chemical Company 1979). Organic 
complexes can hold manganese in solution to higher 
pH levels (Clark et al. 1977). 

As an alternative to precipitation of 
manganese by the high-pH method, strong oxidizers, 
such as permanganate, chlorine dioxide, 
hypochlorite, or ozone can be used. These chemicals 
oxidize manganese to pyrolusite (Mn02) (Evangelou 
1984, Clark et al. 1977). Knocke et al. (1987) 
found that hydrogen pyroxide was not effective for 
oxidizing manganese. 

In the process of removing manganese by the 
high-pH precipitation method, arsenic, chromium, 
copper, lead, mercury, nickel, selenium, and zinc 
should precipitate, principally, as hydroxides 
(Feitkecht and Schindler 1963). These metals may 
also be removed by sorption onto iron precipitates 
(Jones 1986, Mott et al. 1987, Slavek and Pickering 
1986) and as with manganese, complexes and chelates 
may be formed (Martell and Calvin 1952). However, 
hydroxides of these metals may resolubilize if 
solution pH is changed (Peters and Ku 1985). 
Bogner (1983) found that for treated mine 
drainage, the order of extractability from the 
retention pond sediments was Mn > Ni , Zn, Co, > Cu, 
Cr > Fe, Al. Forbes et al. (1976) reported that 
50? of Cd, Zn, Pb, and Cu was desorbed at pH 8, 7, 
5.5, and 5, respectively, from synthetic iron oxide 
(goethite) . 



Regulation of Manganese 

During the process of developing effluent 
limitation guidelines for the coal mining industry, 
the Environmental Protection Agency (EPA) found 
that certain metals, such as arsenic, chromium, 
copper, lead, mercury, nickel, selenium, and zinc, 
were occasionally present in the untreated mine 
water. Water treatment specifically to control 
iron (neutralization and aeration) reduced the 
concentrations of these toxic metals, but not 
consistently enough to provide effective control as 
required by the Clean Water Act of 1972. 

Selection of a specific manganese standard was 
a result of a survey of 314 acid water sources 
(Weideman 1982). In their survey, EPA researchers 
found that the mine water treatment facilities were 
reducing manganese to approximately 2 mg/L, and 
therefore selected that value as the best available 
technology (BAT) effluent limit (U.S. Code of 
Federal Regulations 1985b). In addition, the eight 
toxic metals of concern (arsenic, chromium, copper, 
lead, mercury, nickel, selenium, and zinc) were 
reduced to acceptable levels whenever manganese was 
reduced to 2 mg/L. Therefore, limitations on these 
metals were not promulgated. In their survey of 
acid mine drainage sites, the EPA found average 
manganese concentrations in the untreated mine 
water of 4.9 mg/L at underground mines and 
17.7mg/L at surface mines (Weideman 1982). The 
highest concentration of manganese was 63 mg/L at 
any surveyed site. However, a decade later, our 



84 



work has found many sites at which manganese levels 
exceed those found by the EPA. The average 
manganese concentration from 23 sites with known 
manganese problems was 121 mg/L. Twenty of these 
sites had manganese levels over 63 mg/L (the 
maximum level in the EPA survey) and three sites 
exceeded 200 mg/L (Kleinmann and Watzlaf 1988). 



EXPERIMENTAL PROCEDURES AND RESULTS 

This paper summarizes research performed over 
the past 3 years. In these investigations, five 
mine waters with different chemical compositions 
were used (table 2). This section is divided into 
three subsections: manganese removal, manganese 
resolubilization, and behavior of other metals. In 
the following subsections, several experiments are 
discussed; and although similar procedures were 
used for each test, some modifications were 
necessary to account for the different water 
quality of individual mine waters. Therefore, for 
clarity, the experimental procedures are following 
immediately by the results of that particular test. 

In all tests, pH was measured with an analog 
meter using a combination pH electrode. Acidity 
analysis consisted of a fixed endpoint titration to 
pH 8.3 using a 1N_ sodium hydroxide solution. 
Iron, manganese, aluminum, copper, nickel, zinc, 
and chromium concentrations were determined by 
inductively coupled argon plasma (ICAP) spectros- 
copy. In all tests, immediately after a sample was 
collected for metal analyses, it was passed through 
a 0.45 wm membrane filter and acidified to less 
than pH 1.0 with concentrated hydrochloric acid. 
In all tests, concentrations have been corrected 
for dilution of all added solutions. 



Table 2. — Water quality of mine waters 



Water quality 
parameter 



Concentration of untreated 
AMD* 



II 



III 



IV 



Iron 

Manganese 

PH 

Acidity 

Copper 

Nickel 

Zinc 

Chromium 

Aluminum 



232 


51.3 


7.7 


7300 


39 


126 


87.7 


84.3 


76.5 


44 


5.3 


3-1 


3.1 


2.0 


5.0 


467 


535 


NA 


18475 


NA 


NA** 


NA 


0.27 


22.6 


NA 


NA 


NA 


2.87 


15.8 


NA 


NA 


NA 


4.59 


39.1 


NA 


NA 


NA 


<0.04 


1 .82 


NA 


3.8 


3^.9 


NA 


NA 


NA 



All concentraions in mg/L except pH in pH 
units. Acidity in mg/L as CaCOo. 
NA = Not Analyzed. 



Manganese Removal 

Mine waters I and II were used to compare the 
treatment costs of three chemicals that were 
currently being used by the mining industry to 
remove manganese: sodium hydroxide (NaOH), sodium 
hypochloride (NaOCl), and potassium permanganate 
(KMnOn). These three chemicals represent two 
basic treatment options for manganese; addition of 
enough alkaline material (NaOH or lime) to raise pH 



to approximately 10 or use of a chemical oxidant 
(NaOCl or KMnOij) for direct manganese oxidation. 
Chemical treatment costs were calculated using the 
prices of the chemicals at the time of the study 
(1984) and as delivered to the mine: $0.28/gal and 
$0.80/gal, for bulk quantities of 20-pct-Na0H and 
15-pct-Na0Cl solutions, respectively, and $1.34/lb 
for granular KMnOij. Bulk purchase price of KMnOij 
is about $1.08/lb, but a typical mine would not 
require such large quantities (minimum 20 metric 
tons) . 

For mine water I, 50 liters of raw water was 
collected and treated with a 20-pct solution of 
NaOH to raise pH to 7.5. This water was then 
aerated by pouring it from one container to 
another, causing iron to oxidize and precipitate 
and pH to decrease. The procedure of 
neutralization and aeration was repeated until pH 
stabilized at 7.5. Analysis showed that iron was 
reduced to approximately 4 mg/L and manganese to 
95 mg/L. Addition of alkaline material prior to 
addition of the chemical oxidant is common practice 
since much of the iron can be easily removed and 
since discharge pH must be above 6.0. After 
settling for 30 minutes, twenty 400-ml samples of 
the supernatent water were collected in plastic 
bottles. Six of these bottles received 
incremental amounts of additional NaOH (5-pct 
solution) with pH raised to between 9.2 and 10.5. 
Six bottles received incremental amounts of a KMnOij 
solution (1 pet). Seven bottles received 
incremental amounts of NaOCl (8 pet available 
chlorine) and one bottle received no extra 
chemical addition. After settling for 23 hours at 
room temperature, a sample from each bottle was 
filtered, acidified, and analyzed. All three 
treatments were capable of reducing manganese below 
effluent limits, but sodium hydroxide was 
considerably less expensive than the chemical 
oxidants. Total chemical costs to treat 1 ,000 gal 
of mine water I (including $0.36 for the initial 
NaOH) averaged $1.02, $2.30, and $2.60 for NaOH, 
NaOCl, and KMnOij, respectively (Watzlaf 1985). 

Similar tests were conducted on mine water II. 
A 50-liter sample of water was collected after 
being treated by the existing hydrated lime 
(Ca(0H)2) treatment system. As in the above test, 
the mine water was manually aerated, after which 
the pH stabilized at 8.2. After 30 minutes of 
settling, iron was reduced to 2 mg/L and manganese 
to 70 mg/L. Based on previous lime use records, it 
was estimated that this treatment to pH 8.2 cost 
approximately $0.22 per 1000 gal of water. Twenty- 
seven 450-ml samples were taken from this treated 
water and incremental amounts of NaOH (5-pct 
solution), N0C1 (8 pet available chlorine), or 
KMnOn ( 1 -pet solution) were added to respective 
bottles. These bottles were mixed and left to 
settle for 90 minutes. Samples from each bottle 
were filtered, acidified, and analyzed. Total 
chemical costs per 1 ,000 gal to reduce manganese to 
2 mg/L were $0.53, $0.95, and $1.70 for NaOH, 
NaOCl, and KMnOij, respectively (Kleinmann et al. 
1985). 

In performing these and other tests (Ackman 
and Kleinmann 1985), it was observed that in the 
high-pH precipitation method, manganese was 
removed at lower pH values if the mine water 
contained higher concentrations of iron with 
respect to manganese. The next test used mine 
waters III and IV. Note that the iron to manganese 



85 



concentration ratio in mine water III is 1:11, 
while in mine IV this ratio is 95:1. One liter of 
each mine water was stirred in a large beaker. A 
5-pct solution of NaOH was incrementally added to 
the water. After addition each increment of NaOH, 
pH was measured and a 50-ml sample was taken from 
each beaker. Figure 1 shows the removal of iron 
and manganese as a function of pH. In the high 
iron water, manganese is removed at approximately 
1.5 pH units lower than in the low iron water. 
Presumably, some manganese is removed during the 
precipitation of iron by adsorption on Fe(0H)3 
(Stumm and Morgan 1981). 



was stirred for 15 minutes and 50-ml samples from 
each beaker were filtered and acidified. To each 
beaker incremental amounts of a dilute sulfuric 
acid (H2S0ij) were added to lower pH. At various pH 
levels, 50-ml samples were taken. No differences 
in resolubilization of manganese were found between 
the two high-pH precipitation methods using lime or 
sodium hydroxide (figure 2). Approximately 50 pet 
of the precipitated manganese resolubilized after 
pH was lowered to 3.5. However, no manganese 
resolubilization of manganese occurred for the 
sludge precipitated with the chemical oxidants 
(NaOCl and KMnOij) . 




1 00 



Figure 1.- -Removal of iron and manganese from acid 
mine drainage containing (a) low (7.7 mg/L) 
and (b) high (7,300 mg/L) iron concentrations 
using the high-pH precipitation method 
(sodium hydroxide). 



Manganese Resolubilization 

From the above tests, it was observed that the 
high-pH precipitated manganese would resolubilized 
upon depression of pH. Approximately 800 ml of 
mine water V was placed in four beakers and slowly 
stirred. To the first beaker, a 5-pct solution of 
NaOH was added to raise pH to 10.5 to ensure 
complete manganese removal. To the second beaker, 
a hydrated lime (Ca(0H)2) slurry was added to raise 
pH to 10.5. To the third and fourth beakers, a 
5-pct-Na0H solution was added to raise pH to 7.2. 
The third beaker then received NaOCl treatment 
(2 pet available chlorine) and the fourth received 
a 1 -pet solution of KMnOij. The treated mine water 



<V 


80 


O 

2 


60 


o 
5? 


40 


c 
5 


20 




KEY 
° Hydrated lime 
■ Sodium hydroxide 
♦ Oxidants 



PH 



Figure 2. --Comparison of manganese resolubilization 
from acid mine drainage sludge precipitated 
with sodium hydroxide and hydrated lime 
(high-pH method) and with chemical oxidants 
(Sodium hypochlorite and potassium permanganate) 



Although an elevated iron to manganese ratio 
facilitated manganese removal, it also resulted in 
a less stable sludge. Mine waters III and IV were 
treated with a 5-pct solution of NaOH to pH 10.5. 
Dilute sulfuric acid was added to each mine water 
and samples were taken at various pH levels as in 
the previous test. Figure 3 shows the removal and 
resolubilization of manganese for mine waters III 
and IV, which have iron to manganese ratios of 95:1 
and 1:11, respectively. For the high-iron AMD, the 
removal and resolubilization curve follow the same 
path. Presumably, this indicates that much of the 
manganese was removed by adsorption to ferric 
hydroxide, and subsequently desorbed with 
depression of pH. For the low-iron water, the 
manganese in the sludge was more stable. 

The other factor that influenced the stability 
of manganese in AMD sludge was the age of the 
sludge. Mine water V was treated with NaOH (20-pct 
solution) to pH 10.5 and the water-sludge mixture 
was stored in a loosely covered container. After 
aging for three months, the mixture was stirred and 
H2S0i| was added to lower the pH. Figure *) shows 
that the manganese in AMD sludge becomes more 
stable as it ages. 



86 



lOOp 



1 ' 1 ' — 

(a) Low iron 




mine waters III and IV. The ratios of iron 
concentration to the sum of the concentrations of 
other metals (manganese, copper, nickel, zinc, and 
chromium) were 1:12 and 47:1 for mine waters III 
and IV, respectively. Like manganese, these metals 
were removed at lower pH values in the high-iron 
mine water presumably by adsorption to ferric 
hydroxide. The resolubilization of these metals 
was also influenced by the concentration of iron in 
the mine water (figure 6). These metals were more 
soluble in sludge precipitated from the high iron 
mine water. 

We also investigated the ability to 
permanganate and hypochloride to remove copper, 
nickel, zinc and chromium. NaOH (5-pct solution) 
was added to 1 liter of mine water IV to raise pH 
to 6.0 and was left to settle for 3 days. This 
partially neutralized, water-sludge mixture was 
split into three equal portions and sampled. One 
portion was treated with a 5-pct-NaOH solution to 
pH 10.4, another portion was treated with a 1. 5-pct 
solution of potassium permanganate, and the 
remaining portion treated with an 8-pct available 
chlorine solution of hypochlorite. Analyses showed 
that the initial neutralization to pH 6.0 removed 
all of the chromium and some of the nickel, zinc, 
copper and manganese. Treatment with NaOH removed 
the remaining amounts of these metals. However, 
the oxidants, while effectively removing manganese, 
were unable to remove nickel, zinc, or copper. 



Figure 3. --Removal and resolubilization of 
manganese. Untreated acid mine drainage 
contained (a) low (7.7 mg/L) or (b) high 
(7,3000 mg/L) iron concentrations. Sludge 
was precipitated with the high-pH method 
using sodium hydroxide. 



1 00 




Figure 4. --Comparison of manganese resolubilization 
from freshly precipitated acid mine drainage 
sludge and sludge aged 3 months. Sludge was 
precipitated with the high-pH method using 
sodium hydroxide. 



I OO 




Behavior of Other Metals 

Regulation of manganese was based partly on 
the fact that it acted as a surrogate for removal 
of other potentially more toxic metals. There- 
fore, concurrent with the test on mine waters III 
and IV outlined above, the removal and resolubi- 
lization of other metals were also investigated. 
Figure 5 shows the removal of the other metals from 



Figure 5.--pH required to remove manganese, zinc, 
nickel, copper, and chromium from acid mine 
drainage containing (a) low (7.7 mg/L) and 
(b) high (7,300 mg/L) iron concentrations 
using the high-pH precipitation method 
(sodium hydroxide). 



87 




<£> 



■z. 


RO 


-z. 




< 




UJ 


60 


tr 




7 


40 


w 




o 




tr 


20 


0_ 










1 


1 1 


i ' 

(b) High iron - 


£\ 




KEY 
q Manganese " 
+ Zinc 


— T» a 


t\ \ 


o Nickel 


\ 


H} 


* Copper 

x Chromium - 

=*— 8=9=*} 1— q_ 



7 
PH 



Figure 6. --Resolubilization of manganese, zinc, 
nickel, copper, and chromium from sludge 
precipitated from acid mine drainage 
containing (a) l° w ( 7 - 7 m 9/L) and (b) high 
(7,300 mg/L) iron concentrations. Sludge 
was precipitated with the high-pH method 
using sodium hydroxide. 



DISCUSSION 

In order to comply with effluent limits for 
manganese, mine operators have two main treatment 
options: use of alkaline materials for high-pH 
removal or use of chemical oxidants. The principal 
advantage of the high-pH method is cost. Chemical 
costs for this method are two to three times less 
than for the chemical oxidants. However, at some 
sites, the reduction of manganese to 2 mg/L may 
necessitate discharging water with a pH higher than 
the effluent limit of 9.0. A mine operator in this 
situation is commonly granted a variance that 
permits a high-pH discharge, which may be 
deleterious to the environment (Kleinmann and 
Watzlaf 1986). Another disadvantage of the high-pH 
method is the instability of the sludge it 
produces. Although we found that the manganese in 
the sludge becomes somewhat more stable after three 
months, up to 30 percent can resolubilize at pH 
14.2. The effects of longer aging periods are not 
known. If the sludge is disposed of in or on the 
backfill, any manganese that resolubilizes may need 
to be retreated. The same may be true of 
underground disposal. This cycle of treatment, 
precipitation, sludge disposal, resolubilization, 
and retreatment may be repeated indefinitely at 
great expense to the operator. In fact, the high 
manganese concentrations currently found in some 
mine drainage may be due to sludge resolubi- 
lization. 



The use of chemical oxidants produces a stable 
manganese precipitate, however, treatment costs are 
greatly increased. These chemicals also require 
very accurate control of dose rate. In order for 
hypochlorite to be effective, enough chemical must 
be added to produce a chlorine residual. With 
adequate detention time and proper chemical 
dosage, this chlorine residual will dissipate 
before discharge. However, overtreatment and pond 
short-circuiting could result in discharge of high 
residual chlorine levels and damage to the 
environment. Similarly, the dosage of permanganate 
needs precise control. Too much permanganate will 
result in excess MnOi) - , which will place the 
effluent in violation of the manganese limit. 
Also, although very effective at removing 
manganese, these chemicals do not remove nickel, 
zinc, or copper. 

The problem of resolubilization of manganese 
and other metals from acid mine drainage sludge is 
potentially very significant. Millions of tons of 
AMD sludge have been and are being disposed of with 
little or no concern to potential metal 
resolubilization. The resolubilization of this 
sludge could cause mining companies significant 
increases in treatment costs or, if released to the 
environment, substantial environmental damage. 
Additional research is needed to determine the 
effects of long-term aging on AMD sludge in the 
treatment pond and after disposal. Other factors 
involved in removal and resolubilization need to be 
delineated in order to evaluate current treatment 
and disposal practices. 



CONCLUSIONS 

Tests using five different mine waters 
investigated the removal and resolubilization of 
manganese as well as other metals from AMD using 
various chemical techniques. The major 
conclusions of this study are: 

1 . When sodium hydroxide or lime was used, 
minimum pH values necessary to reduce 
manganese below 2.0 mg/L ranged from 8.4 to 
10.2. 

2. The pH necessary to remove manganese from 
mine water with an iron:manganese ratio of 
95:1 was approximately 1.5 pH units less than 
from a mine water with an iron:manganese ratio 
of 1:11. 

3. In the high-iron mine water (7,300 mg/L), 50J 
of manganese, nickel, zinc, and copper 
resolubilized at pH 6.6, 5.8, 5.6, and 3-5, 
respectively. 

4. In the low-iron mine water (7.7 mg/L), 50? of 
manganese, nickel, zinc, and copper 
resolubilized at pH 5.6, 4.5, 4.7, and 4.3, 
respectively. 

5. Manganese was stable to pH 3-5 in the sludges 
generated with permanganate or hypochlorite. 

6. The high-pH method using sodium hydroxide 
significantly reduced concentrations of 
nickel, zinc, and copper; however, 
permanganate and hypochlorite did not. 



88 



LITERATURE CITED 



Ackman, T. E., and R. L. P. Kleinmann. 1985. In- 
line aeration and treatment of acid mine 
drainage: performance and preliminary design 
criteria. In Control of Acid Mine Drainage, 
Bureau of Mines IC 9027, pp. 53-61. 

Bogner, J. E. 1983. Prediction of extractable 

metals in retention pond sediments in surface 
coal mines. Environ. Geol. 4:223-238. 

Clarke, J. W. , W. Viessman, and M. J. Hammer. 
1977. Water supply and pollution control, 
pp. 444-495. Harper and Row, New York, NY. 

Collins, J. F., and S. W. Buol. 1970. Effects of 
fluctuations in the Eh-pH environment on iron 
and/or manganese equilibria. Soil Science. 
1 10:111-118. 

Diem, D., and W. Stumm. 1984. Is dissolved Mn 2+ 
being oxidized by O2 in absence of Mn- 
bacteria or surface catalysts? Geochimica et 
Cosmochimica Acta 48:1571-1573- 

Evangelou, V. P. 1984. Controlling iron and 
manganese in sediment ponds. Reclamation 
News and Views. 2:1-6. 

Feitknecht, W. , and P. Schindler. 1963. 

Solubility constants of metal oxides, metal 
hydroxides and metal hydroxide salts in 
aqueous solution. Pure and Applied Chemistry. 
6:130-199. 

Forbes, E. A., A. M. Posner, and J. P. Quirk. 

1976. The specific adsorption of divalent 
Cd, Co, Cu, Pb, and Zn on goethite. Journal 
of Soil Science. 27:154-166. 

Hem, J. D. 1981. Rates of manganese oxidation in 
aqueous systems. Geochimica et Cosmochemica 
Acta. 45:1369-1374. 

Holland, C.T., J. L. Corsano, and D. J. Ladish. 

1968. Factors in the design of an acid mine 
drainage treatment plant. In proceedings, 
Second Symposium on Coal Mine Drainage 
Research; Mellon Institute; Pittsburgh, PA; 
May, 1968; pp. 274-290. 

Jones, K. C. 1986. The distribution and 

partitioning of silver and other heavy metals 
in sediments associated with an acid mine 
drainage stream. Environmental Pollution 
(Series B) . 12:249-263. 

Kleinmann, R. L. P., and G. R. Watzlaf. Should the 
discharge standards for manganese be 
reexamined?_ In proceedings, Symposium on 
Mining, Hydrology, Sedimentology, and 
Reclamation; University of Kentucky; 
Lexington, Kentucky; December 8-11, 1986; 
pp. 173-179. 

Kleinmann, R. L. P., and G. R. Watzlaf. 1988. 

Should the effluent limits for manganese be 
modified? In proceedings, Mine Drainage and 
Surface Mine Reclamation Conference; American 
Society for Surface Mining and Reclamation and 
U. S. Dept. of the Interior; Pittsburgh, PA; 
April 17-22, 1988. (In Press). 



Knocke ,W. R., R. C. Hoehn, and R. L. Sinsabaugh. 
1987. Using alternative oxidants to remove 
dissolved manganese from waters laden with 
organics. Journal AWWA. March 1987, pp. 75- 
79. 

Martell, A. E., and M. Calvin. 1952. Chemistry of 
the metal chelate compounds. Prentice-Hall, 
New York, NY. 61 3 pp. 

Morgan, J. J. 1967. Chemical equilibria and 
kinetic properties of manganese in natural 
waters, pp. 561-624. in Principles and 
applications ofwater chemistry. S. D. Faust 
and J. V. Hunter (Eds.). John Wiley and 
Sons, Inc., New York, NY. 

Morgan, J. J., and W. Stumm. 1964. Colloid- 
chemical properties of manganese dioxide. 
Journal of Colloid Science. 19:347-359. 

Mott, H. V., K. E. Hartz, and D. R. Yonge. 1987. 
Metal precipitation in two landfill leachates. 
Journal of Environmental Engineering. 
113:476-485. 

Murray, J. M. , J. G. Dillard, R. Giovanoli, H. 
Moers, and W. Stumm. 1985. Oxidation of 
Mn(II): initial mineralogy, oxidation state 
and ageing. Geochemica et Cosmochimica Acta 
45:463-470. 



Nalco Chemical Company. 1979. 
handbook. Chapter 6; pp. 
New York, NY. 



The NALCO water 
15-16. McGraw-Hill, 



Nicholas, G. D., and E. G. Foree. 1979. Chemical 
treatment of mine drainage for removal of 
manganese to permissible limits. In 
proceedings, Symposium on Surface Mining 
Hydrology, Sedimentology, and Reclamation; 
University of Kentucky; Lexington, Kentucky; 
December 4-7, 1979; pp. 181-187. 

Owens, L. V. 1963. Iron and manganese removal by 
split flow treatment. Water and Sewage Works. 
110:R76-R87. 

Peters, R. W. , and Y. Ku. 1985. Batch 

precipitation studies for heavy metal removal 
by sulfide precipitation. AIChE Symposium 
Series. 81:9-27. 

Sienko, M. J. and R. A. Plane. 1966. Chemistry: 
principles and properties, pp. 418-421 . 
McGraw-Hill, New York, NY. 

Slavek, J. and W. F. Pickering. 1986. Extraction 
of metal ions sorbed on hydrous oxides of iron 
(III). Water, Air, and Soil Pollution. 
28:151-162. 

Stumm, W. and R. Giovanoli. 1976. On the nature 
of particulate Mn in simulated lakewaters. 
Chemica 30:423-425. 



Stumm, W. and J. 
Chemistry. 
780 pp. 



J. Morgan. 1981. Aquatic 
Wiley-Interscience, New York, NY. 



89 



U.S. Code of Federal Regulations. 1985a. Title 
30--Mineral Resources; Chapter VII--Office of 
Surface Mining Reclamation and Enforcement, 
Department of Interior; Subchapter B — Initial 
Program Regulations; Part 715 — General 
Performance Standards. July 1, 1985. 

U.S. Code of Federal Regulations. 1985b. Title 

40 — Protection of the Environment; Chapter 1 — 
Environmental Protection Agency; part 434 — 
Coal Mining Point Source Category; Subpart C — 
Acid or Ferruginous Mine Drainage; July 1, 
1985. 

Watzlaf, G. R. 1985. Comparative tests to remove 
manganese from acid mine drainage. In Control 
of Acid Mine Drainage, Bureau of Mine3 IC 
9027, pp. 41-17. 

Weideman, A. 1982. Development document for 

effluent limitations, guidelines 'and standards 
for the coal mining point source category — 
final. 654 pp. EPA, Washington, DC. 



90 



TREATMENTS TO COMBAT PYRITE OXIDATION 
IN COAL MINE SPOIL 



C.A. Backes, I.D. Pulford, and H.J. Duncan 



Abstract. Oxidation of pyrite and its inhibition by 
various waste materials were studied using pyritic 
spoil from Central Scotland, with the aim of develop- 
ing procedures to prevent reacidif ication after the 
initial treatment. It had previously been shown 
that the rate of pyrite oxidation could be decreased 
by complexing or precipitating the iron species in- 
volved. Analysis of leachate from a set of lysi- 
meters showed that the rate of pyrite oxidation in- 
creased with the onset of warmer weather, but was 
greatly reduced during winter. A controlled tempe- 
rature incubation experiment investigated the inhi- 
bition of pyrite oxidation by chicken manure or coni- 



fer bark at 0, 8, 18 and 30 C 



There was no signifi- 



cant pyrite oxidation in the limed spoil. In the un- 

limed spoil more oxidation occurred at 18 and 30 C 

o 
than at and 8 C. Conifer bark reduced the rate of 

pyrite oxidation at all temperatures, and at 18 C it 
reduced it to that found at the lower temperatures. 
Chicken manure completely inhibited acid and iron pro- 
duction at all temperatures. A small field trial 
investigated the use of chicken manure, conifer bark, 
sewage sludge, and oil shale waste in conjunction with 
lime to inhibit the rate of pyrite oxidation. The 
plots were seeded with grass, and vegetation yield 
and spoil characteristics measured. These plots are 
being continued, to study the long-term effects of 
the amendments on pyrite oxidation. 



INTRODUCTION 

Pyrite is present in many of the 
colliery spoil heaps (bings) in Central 
Scotland, and is the major obstacle to 
their reclamation. Acidic spoil, resul- 
ting from the oxidation of the pyrite, is 
traditionally neutralised by the addition 
of lime. However, after a few years bings 
suffer from acid regeneration, as the lime 
is leached out and more pyrite oxidises. 
The reacidif ication tends to occur in loca- 
lised areas as the distribution of pyrite 



C.A. Backes is a Post-Doctoral Research 
Assistant, I.D. Pulford a Lecturer and H.J. 
Duncan a Reader, Agricultural Chemistry 
Section, Chemistry Department, University 
of Glasgow, Glasgow G12 8QQ, Scotland, U.K. 



in a bing if; seldom homogeneous (Backes 
1984). These acidified areas tend to 
spread, killing any established vegetation. 

Many bings receive only the initial 
reclamation treatment and, even when large 
quantities of lime are added in an attempt 
to account for potential acidity, either 
in one or a number of applications, the 
bings still tend to suffer from acid 
regeneration. Doubleday (1974), for 
example, found that four applications of 
limestone totalling 100 t/ha over a two- 
year period were inadequate to neutralise 
the acidity of one spoil; Costigan et al. 
(1981) found spoils limed at either 50 or 
100 t/ha which reacidified after one or two 
years. Some bings do receive aftercare, 



91 



especially if they are to be used for agri- 
cultural purposes. In these cases it may 
be possible to apply additional lime to 
small areas of the bing as they reacidify, 
to neutralise the acid and inhibit the 
spread of pyrite oxidation. However, it 
would be preferable if there were one 
initial treatment, which not only neutra- 
lised the acid already present in the bing, 
but also prevented further reacidif ication 
of the spoil. This paper proposes such 
treatments . 

There are two main pathways for pyrite 
oxidation, either using oxygen or ferric 
ions as the oxidising agent. Above about 
pH 4.5 oxidiation will only occur by the 
oxygen pathway. This is a relatively slow 
process and does not produce large quanti- 
ties of acid. However, if the pH falls 
to below 4.5 then ferric ions act as an 
oxidising agent on pyrite. The ferric 
ions are produced by the bacterially cata- 
lysed oxidation of ferrous ions, which are 
liberated from the pyrite. This leads to 
a cyclic system, in which ferrous ions are 
bacterially oxidised to ferric ions, which 
are then reduced by their action on pyrite 
back to ferrous, simultaneously releasing 
more ferrous ions from the pyrite. This 
is a much faster reaction than oxidation 
by oxygen, leading to a rapid release and 
build-up of acid. Backes, Pulford and 
Duncan (1986, 1987) showed that the ferric 
ion oxidation pathway could be inhibited by 
the use of specific chemicals to interfere 
with the cycling of iron. Phosphates and 
silicates were used to precipitate the iron, 
whilst citrates were used to complex it. 
Use of a bactericide stopped the production 
of ferric ions by the action of Thiobaci llus 
f er r ooxidans . These studies, whic 



for existing acidity plus 25 t/ha for po- 
tential acidity. Large areas of the bing 
have since suffered from severe acid re- 
generation. 



carried out in suspensions of pyrit 
waste, showed that the treatments r 
ted the oxidation pathway to that b 
gen alone, so decreasing the rate a 
acid was released. Waste material 
also used to inhibit the oxidation 
rite by ferric ions. Chicken manu 
chopped conifer bark both complexed 
iron, while pulverised fuel ash (a 
of silicate) precipitated it. The 
reported here extends these laborat 
studies on waste materials into con 
pot and field studies. 



h were 
ic coal 
estric- 
y oxy- 
t which 
s were 
of py- 
re and 

the 
source 

work 
ory 
trolled 



METHODS 



Coll 

of G 

fiel 

The 

prev 

to 2 

Colb 

riti 

fiel 

hete 

disc 

was 

lime 



Pyr 
i er y 
reat 
d wo 
pyri 
ious 
.92% 
ourn 
c ma 
d wh 
roge 
usse 
orig 

req 



itic 

, Wes 
Brit 
rk wa 
tic c 
ly me 
, usi 
(197 
ter ia 
ich c 
neity 
d by 
inall 
uirem 



spoi 
t Lo 
ain 
s co 
onte 
asur 
ng t 
9). 
1 co 
onta 

of 
Back 
y re 
ent 



1 was 
thian 
ref NT 
nducte 
nt of 
ed as 
he met 
Howe 
uld ea 
ined u 
pyr it i 
es (19 
c laime 
was 2 5 



obta 
(Ord 

002 
d on 
the 
rang 
hod 
ver , 
sily 
p to 
c sp 
84) . 
d in 

t/h 



ined 
inan 
609) 

thi 
spoi 
ing 
of D 

lum 

be 

75% 
oil 
B 

197 
a to 



from Baads 
ce Survey 
, and all 
s si te . 
1 had been 
from 0.07 
acey and 
ps of py- 
found in the 

FeS„. The 
has Been 
aads bing 
9, when the 

account 



Field Lysimeters 



the fie 
pyr it ic 
as the 
volume 
sur r oun 
lected 
Whatman 
analyse 
tration 
phthale 
iron by 
The wee 
mum air 
rain wh 
was dis 
very he 



sen 
Id i 

spo 
cont 
of 
ding 
week 
's N 
d fo 

wit 
in e 

ato 
kly 

tem 
ich 
card 
avy 



es of 
n Febr 
il and 
rols, 
.Olni 

spoil 
ly and 
o. 1 f 
r P H, 
h sodi 
nd poi 
mic ab 
rainf a 
peratu 
had no 
ed (th 
rainf a 



ly s 

uar 

tw 

E 

and 

f i 
ilt 
tit 
urn 
nt , 
sor 

11, 

res 
t p 
is 
11) 



imet 

y 19 

o wi 

ach 

was 

The 

Iter 

er p 

rata 

hydr 

and 

ptio 

and 

wer 

enet 

occu 



er s 
87, 
th n 
lysi 

imb 

lea 
ed t 
aper 
ble 
oxid 

tot 
n sp 

max 
e al 
rate 
rred 



was set up in 
five containing 
onpyritic spoil 
meter had a 
edded into the 
chate was col- 
hrough 

It was 
acidity by ti- 
e to phenol- 
al soluble 
ectro photometry, 
imum and mini- 
so noted. Any 
d the spoil 
only after 



Controlled Temperature Incubation Study 

Bags of pyritic spoil, previously air- 
dried and sieved through a 4 mm sieve, 
were placed at four different temperatures; 
0, 8, 18 and 30°C. After a few days, 
when the spoil had equilibrated, it was 
used to prepare a series of pots for each 
temperature. To four replicate pots of 
500 g spoil, the following were added: no 
addition (control); 15 g lime (equivalent 
to 15 t/ha, the nominal lime requirement); 60 g 
chopped conifer bark (60 t/ha); 50 g chicken manure 
(50 t/ha); conifer bark plus lime at the above rates; 
chicken manure plus lime at the above rates. The 
spoil and amendment were thoroughly mixed 
before being placed in a pot. All sets 
of pots were returned to their original 
temperature and watered regularly to main- 
tain them at near field capacity. At 
intervals, all the pots were leached with 
100 cm distilled water, and the leachate 
filtered and analysed as for the lysimeter 
study . 

Field Study 



Small plots (2 m by 2 m) were pre- 


pared on the area of the bing that had suf- 


fered the most severe acid regeneration, 


where all the vegetation had been 


destroyed 


All plots, apart from the control 


s , were 


treated with lime at 15 t/ha and 


one of the 


following: chicken manure at 25 


t/ha , 


sewage sludge at 20 t/ha, conifer 


bark at 


60 t/ha, conifer bark at 30 t/ha, 


oil s h a le 


waste at 50 t/ha. One plot was 


left as 


lime only. The amendments were 


incorpo- 


rated into the top 15 cm of spoil 


All 


treatments were replicated three 


times in 


a randomised block design. The 


area was 


then fertilised (300 kg/ha of 15: 


10:10 NPK 


compound fertilizer) and seeded with an 


agricultural grass mix (100 kg/ha 


) . The 


plots were sown in August 1986, 


and at 


the beginning of the next spring 


they were 


r ef er tilized . The grass was harvested in 


1987, and its yield measured. The spoil 



92 



of each plot was sampled in duplicate and 
analysed for pH (1:2.5 spoil: water) and 
water-extractable iron. 

RESULTS AND DISCUSSION 

Lysimeter Study 

The rate of acid and iron production 
in the lysimeter study reflected the rate 
of pyrite oxidation (figs 1, 2). The 
cumulative rate was obtained by multi- 
plying the weekly rainfall by the concent- 
ration of acid or iron and then summing 
over the appropriate weeks. Although the 
spoil for all the pyritic lysimeters was 
obtained from a small area, the quantity 
of acid and iron produced varied conside- 
rably, reflecting the heterogeneity of the 
pyritic content of the spoil. Initially 
the rate of oxidation was very slow, but 
between weeks 14 and 16 (at the end of May) 
the rate increased dramatically. The 
limits of the change of rate were shown in 
lysimeters 1 and 5; the rate of acid pro- 
duction per lysimeter changed from 7 ug 
H per week to 136 ug H per week in lysi- 
meter 1 and from 11 to 60 ug H per week 
in lysimeter 5. The rate of iron pro- 
duction changed from 75 Ug Fe per week to 
1900 ug Fe per week for lysimeter 1 and 
from 54 Ug Fe to 500 Ug Fe per week in 
lysimeter 5. This increase in rate 
occurred in all the lysimeters (apart 
from the controls) over the same two-week 
period; the rate then stayed approximately 
constant for each lysimeter until week 32, 
when the rate decreased in all lysimeters. 
This increase and decrease in rate closely 
corresponded to the measured air tempera- 
ture (fig 3). At week 16, the weekly 

minimum temperature began to rise above 

o 
C, whereas the maximum temperature had 

begun to rise about week 9. The rate of 
acid production would probably correlate 
most closely to the average temperature, 
rather than the maximum or minimum. The 
approximately constant rate after week 16 
would suggest that there was a rate limi- 
ting mechanism operating during the summer 
periad. The solubility constant (Ksp(Fe) 
(UH^ R for ferric hydroxide ranges from 4 x 
10 to 2.7 x 10 , depending on which 
form of hydroxide precipitates. During 
the summer period the leachate was super- 
saturated with respect to these oxides, 
suggesting that one of these forms was con- 
trolling the concentration of iron in solu- 
tion, and thus limiting the rate of pyrite 
oxidation. Other iron compounds can pre- 
cipitate, such as jarosite and a large 
number of intermediate iron products, whose 
solubility would be greater than that of 
the hydroxides. It would depend on the 
exact local conditions as to which iron 
compound precipitated (Nordstrom 1982). 
In the colder period, the leachate was 
supersaturated with respect to the stable 
iron oxides, but not so with respect to the 
more amorphous oxides. As it would be 
expected that the amorphous oxides would 
precipitate in this situation, it seems 



X 
o> 

E, 

c 
o 

o 

3 

■o 
o 

k. 

a. 
•a 
"5 

< 



Fit 



1 - 



+ lysimeter 1 

♦ lysimeter 2 
x lysimeter 3 

♦ lysimeter 4 
■ lysimeter 5 
B controls 



++ 



+ 



♦ XX 



X* 



x« 



♦ « 



• * - 

a ■ 



X <>♦ 

• ■■ 






**■■ 



q -| ftA&Pflfl-B3 — B — Hnroooo? nn ip P p nnnnoo 

10 20 30 40 

Week 

1. Cumulative acid production in lysimeters. 
40 



♦ ♦ 







+ 


lysimeter 1 






♦ 


lysimeter 2 


£ 


30- 


X 


lysimeter 3 


o> 




• 


lysimeter 4 


E 




■ 


lysimeter 5 


c 




a 


controls 


o 


20- 






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3 








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o 








k. 








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8* 
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+ * 

4^ — P B i 00009 — o-a 



go 9 n oaqoa 1 

10 20 30 40 

Week 

Fig 2. Cumulative iron production in lysimeters, 





3- 


• 


• 


♦ ♦ 






-25 












• ♦ 


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I 






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Fig 3. Relationship of acid production to 
air temperature ( lysimeter 1, 
maximum T, minimum T). 



93 



that in the cold period precipitation of 


ferric ions was not occurring and thus not 


rate limiting. Apart from the precipi- 


tation of an iron hydroxide, there are th ree 


other factors which can limit the rate of 


pyrite oxidation by ferric ions in the 


field. The rate limiting step in the oxi- 


dation pathway can be either the bacterial 


oxidation of ferrous to ferric ions, or the 


actual ferric ion oxidation of pyrite, or 


the surface area of the pyrite available 


for reaction with the ferric ions (Garrels 


and Thompson 1960, Bloomfield 1972, van 


Breemen 1973). As the rate of oxidation 


was much greater in the summer period than 


in the colder period, it would suggest that 


the last two mechanisms would not be rate 


limiting in the colder period, otherwise 


they would have decreased the summer rate. 


However, it is very probable that the rate 


of ferrous ion oxidation by Thiobacillus 


ferrooxidans was decreased due to the fall 


in temperature, or it is possible that this 


process was completely absent, restricting 


the oxidation of the pyrite to that by oxy- 


gen alone, which is a much slower process. 



Controlled Temperature Incubation Study 

The limed pots at all temperatures 
maintained a pH above 5.5. This high pH 

would restrict the rate of pyrite oxi- 
dation to that by oxygen alone, and there- 
fore no conclusions can yet be drawn as to 
the effect of amendment and lime on the 
rate of pyrite oxidction from this study. 



oxid 
for 
same 
rate 

16 u 

H + p 

of i 

an 

70 u 

for 

call 

for 

mark 

ture 

stud 



In 
atio 
the 

at 

of 
g H + 
er d 
r on 
d 8° 
8 Fe 
acid 
y as 
acid 
ed d 
s as 

y • 



the 

n va 

firs 

all 

acid 

per 

ay a 

prod 

C, 4 

per 

pro 

the 

and 

if fe 

had 



unlime 

ried w 

t 40 d 

temper 

produ 

day a 

t an 

uction 

Ug F 

day a 

ductio 

overa 

iron) 

rence 

occur 



d po 

ith 
ays, 
atur 
ctio 
t 18 
d 8° 
was 
e pe 
t 18 
n ar 
11 t 

in r 
red 



ts t 
temp 

it 
es . 
n wa 

and 
C (f 

10 
r da 
°C. 
e pr 
rend 
This 
ate 
in t 



he ra 
eratu 
was n 
Aft 
s app 

30°C 
ig 4) 
Ug Fe 
y at 

(On 
esent 
s wer 

was 
betwe 
he ly 



te of 
re , a 
early 
er th 
r oxia 

and 
T 

per 
30°C, 
ly re 
ed gr 
e the 
not s 
en t e 
simet 



pyrite 
lthough 

the 
is the 
mtely 
6 Ug 
he rate 
day at 

and 
suits 
aphi- 

same 
uch a 
mpera- 
er 



The conifer bark reduced the rate of 
pyrite oxidation at all temperatures, and 
at 18 C it reduced it to the rate at the 
lower temperatures (figs 5, 6, 7). The 
chicken manure completely inhibited the 
rate of acid and iron production at all 
temperatures (figs 6, 7, 8). 

After day 155, the pots at and 8 C 
were moved to 8 and 18 C respectively to 
simulate the rise in temperature occurring 
in the field. The pots at 18 C were not moved 
as it is only rarely that the spoil would 
maintain a temperature above this in the 
field in Scotland. The pots at 30 C were 
indicative of the maximum acid and iron 
production, although more iron was actually 
produced in the 18 C pots (lime-only pots). 
There was no discernible change in the rate 
of pyrite oxidation in the and 8 C pots 



when the temperature was raised to 


8 and 


18 C respectively. 




Conifer bark and chicken manure both 


complex iron (Backes 1984, Vaughan 


, Wheatley 


and Ord 1984), either by chelating 


it on to 


the solid material or by forming a 


stable 


complex in solution. Both these amendments 


decreased the rate of pyrite oxidation at 


all temperatures, presumably by complexing 


the iron and inhibiting the ferric 


ion oxi- 


dation of pyrite. The amendments 


could 


have been complexing the products 


of the 


pyrite oxidation, reducing the measurable 


amounts of acid and iron, leading 


to the 


false conclusion that the actual rate of 


oxidation was decreased. However, 


in view 


of the lysimeter results, it would 


seem 


that at least at the higher temperatures 


they were actually inhibiting the 


ferric 


ion oxidation of pyrite. 





X 
en 

E 



o 

3 
"O 
O 
k. 
Q. 
■D 
O 
< 



4- 
















D 


T=0 








■ 


3- 


X 

♦ 


T=8 
T=18 






• 


♦ 




■ 


T=30 




i 






2- 






• 












■ 








■ 


1 - 




* 
■ 

♦ 

- x » 


8 


a 
» 


a 

X 






■ 


t a 










o- 


I ° 


a 


1 




— i ■ 


— i ■ 1 



50 



100 



150 



200 



250 



Day 



Fig. 4 Cumulative acid production in control 
pots at all temperatures. 



2.5 



I 
ra 
E, 

c 
o 

t3 

■D 

O 



u 

< 



2.0 



1.5 



1.0 



0.5 



0.0 



D 


T=0 








■ 


X 


T=8 










• 


T=18 










■ 


T=30 






■ 




- 






■ 






■ 




■ 








- 


■ 






D 


D 


■ 
I 8 


■ 

a 
g » * 

1 ' 


D 
D 


D 
ft 


• 
X 


t 

- 1 ■ 1 



50 100 150 

Day 



200 250 



Fig. 



5 Cumulative acid production in pots 
amended with conifer bark at all 
temperatures. 



94 



<? 3- 



t> 2 



Q. 

" 1 -I 

< 1 



D control 
x bark 
• manure 



g » 



' 1 ' 1 1 1 1 1 

50 100 150 200 250 



Fit 



Day 



. 6 Cumulative acid production in unlimed 

,0,. 



pots at 18 C. 



I 
oi 
E 

c 
g 

o 

3 

■o 
o 



4- 


D 


control 






D 


3- 


X 




bark 
manure 


D 


a 


X 


2- 




a 

D 

□ x 


X 


X 




1 ^ 


a 


X 
X 










° X 










n- 


8 J • 


• • ♦ 





♦ 


— 1 ' 1 



Fig 



50 100 150 200 250 

Day 



7. Cumulative acid production in unlimed 
pots at 30 C. 



X 
E 



o 

3 

■o 
o 

Q. 

■a 



a " ♦ x 

■ x 
* 



° I 



D • 



° T=0 

* T=8 

• T-18 
■ T=30 



50 



100 



150 



200 



250 



Day 



Fig 8. 



Cumulative acid production in pots 
amended with chicken manure at all 
temperatures. 



Table 1 



Treatment 



Vegetation yields on limed 
plots amended with various 
organic treatments, 1 year 
after establishment. 



Chicken manure 
60 t/ha conifer bark 
Sewage sludge 
30 t/ha conifer bark 
Lime only 
Oil shale 



Mean yield of 
vegetation (and s.d.) 

(g/O 

374 (87) 

321 (105) 

320 (114) 

286 (99) 

262 (56) 

233 (28) 



Field Study 



After one year 


the p 


H of spoil 


in all the plots, a 


part f 


rom the unamended 


controls, was above 


6.5. 


There was a 


large variance in the yie 


Ids from each 


treatment, which means that statistically 


the differences in 


yields 


were not signi- 


f icant ( table 1 ) . 


However, the general 


trend was that all 


the amended plots, 


except for those treated 


with oil shale, 


gave greater yields 


than 


those treated wit 


lime alone. In th 


e case 


of the sewage 


sludge and chicken 


manure 


plots, this may 


have been due to an 


added 


nutrient effect, 


but this would not 


be the 


case for the 


conifer bark plots. 


Therefore there was 


probably some benef 


icial 


effect occurring 


in terms of inhibition of 


pyrite oxidation 


The oil shale waste 


was rather an inert 


material, and may require 


attack by acid 


before it releases 


silica 


This field 


trial is being continued, 


so that when the 


added lime has been 


neutralised or leached 


out the effectiveness of 


the amendments at 


inhibiting pyrite oxidation in the field 


can be measured. 







CONCLUSIONS 



Th 
point t 
cation 
amendme 
acidic 
incor po 
otherwi 
for the 
This wo 
of pyri 
pH , as 
and Thi 
operati 
would b 
rite ex 
neutral 
leached 
would t 
plexing 
thus re 
to that 
prevent 
spoil , 
establi 
source 
A conti 
the spo 
nal ame 



e pr 

o wa 
of c 
nts 
bing 
rati 
se t 
est 
uld 
te b 
f er r 
obac 



elim 
y s o 
olli 
woul 

in 
on i 
he p 
abli 
lead 
eing 
ic h 
illu 



inary r 
f inhib 
er y spo 
d have 
con junc 
nto the 
H would 
shment 
initia 
inhibi 
ydroxid 
s f er ro 



esul 
itin 
il. 
to b 
t ion 

plo 

be 
of v 
lly 
ted 
e wo 
oxid 



ts of 
g the 
Any 
e add 
with 
ugh 1 
too 1 
egeta 
to th 
due t 
uld p 
ans w 



ve . 
e re 
idat 
i se 

out 
hen 

any 
stri 

by 

the 
alio 
shed 
of o 
nual 
il w 
ndme 



In 
duce 
ion 
the 

of 
star 

iro 
ctin 
oxyg 

rap 
wing 
, wh 
rgan 

sup 
ould 
nt h 



time t 

d , as t 
produce 
lime, a 
the spo 
t to be 
n that 
g the o 
en alon 
id reac 

a v ege 
ich its 
ic matt 
ply of 

ensure 
ad been 



he e 
he s 
d ac 
nd a 
il. 

ef f 
came 
xida 
e . 

idif 
tati 
elf 
er f 
orga 
tha 
com 



f f ect 

low r 
id wh 
s 1 im 
The 
ec t i v 

into 
tion 

This 
icati 
on co 
would 
or th 
nic m 
t one 
plete 



thi 

rea 

of 
ed t 

lim 
ay er 
ow t 
tion 
e ox 
o th 
reci 
ould 

of 
ate 
ich 
e wa 

ame 
e , b 

sol 
of p 

sho 
on o 
ver 

act 
e sp 
at te 
e th 
ly d 



s study 
cidif i- 
these 
o an 
e , by 
, as 
o allow 

idation 
e high 
pitate 

be in- 
the lime 
of py- 
woul d 
s 

ndments 
y com- 
ut ion , 
yrite 
uld 
f the 
to be 

as a 
oil. 
r to 

e origi- 
egr aded 



95 



there would still be an iron complexing Van Breemen, N. 1973. Soil forming 
source available. The build-up of a vege- processes in acid sulphate soils, 
tation cover would also restrict the infil- pp 66-128. In Acid sulphate soils, 
tration of air into the spoil, thus decrea- H. Dost (Ed.). International Inst, for 
sing the rate of pyrite oxidation further. Land Reclamation and Improvement. Pub, 

If the amendments are added in the colder 18. 

period of the year, it may help to ensure 

that all iron produced is immediately com- Vaughan, D., R.E. Wheatley and B.G. Ord 
plexed, thus precluding the onset of rapid 1984. Removal of ferrous ion from 

reacidif ication of the spoil. field drainage waters by conifer bark 



J. Soil Sci. 35:149-153. 



ACKNOWLEDGEMENTS 

We would like to acknowledge the Scottish 
Development Agency for financial support, 
and Lothian Regional Council for access to 
Baads bing . 

LITERATURE CITED 

Backes, C.A. 1984. The oxidation of pyrite 
and its environmental consequences. 
Ph.D. Thesis, University of Glasgow. 
281 pp. 

Backes, C.A., I.D. Pulford and H.J. Duncan 

1986. Studies on the oxidation of 
pyrite in colliery spoil. I. The 
oxidation pathway and inhibition of 
the ferrous-ferric oxidation. Reclam. 
Reveg. Res. 4:279-291. 

Backes, C.A., I.D. Pulford and H.J. Duncan 

1987. Studies on the oxidation of 
pyrite in colliery spoil. II. 
Inhibition of the oxidation by amend- 
ment treatments. Reclam. Reveg. Res. 
6:1-11. 

Bloomfield, C. 1972. The oxidation of 
iron sulphides in soils in relation 
to the formation of acid sulphate 
soils, and ochre deposits in field 
drains. J. Soil Sci. 23:1-16. 

Costigan, P. A., A.D. Bradshaw and R. P. 
Gemmell 1981. The reclamation of 
acidic colliery spoil : 1. Acid 
production potential. J. Appl. Ecol. 
18:865-878. 

Dacey, P.W. and P. Colbourn 1979. An 

assessment of methods for the 

determination of iron pyrites in coal 

mine spoil. Reclam. Reveg. Res. 
2: 113-121. 

Doubleday, G.P. 1974. The reclamation of 
land after coal mining. Outlook on 
Agric. 8:156-162. 

Garrels, R.M. and M.E. Thompson 1960. 

Oxidation of pyrite by iron sulfate 
solutions. Am. J. Sci. 258A:57-67. 

Nordstrom, K.K. 1982. Aqueous pyrite 

oxidation and the consequent formation 
of secondary iron minerals, pp. 37-57 
In Acid sulfate weathering. J. A. 
Kittrick, D.S. Fanning and L.R. 
Hossner (Eds.). Soil Sci. Soc. Am., 
Special Pub. 10. 



96 



A GROWTH INHIBITION MODEL FOR THIOBACILLUS FERROOXIDANS J 



Kay L. Shuttleworth and Richard F. Unz' 



Abstract.-- Acidophilic strains of the genus 
Thiobacillus play a major role in the genesis of 
acidic mine drainage. The development of these 
chemolithotrophs has been shown to be adversely 
affected by the presence of certain organic acids in 
laboratory cultures. Conversely, acidophilic 
heterotrophic bacteria, e.g., members of the genus 
Acidiphilium , are capable of metabolizing a variety 
of organic acids. A theoretical model was developed 
which mathematically depicts a possible interaction 
between these two physiologically distinct types of 
microorganisms. The model is based on the Monod 
kinetics model modified to account for the effects of 
exogenous inhibitors on the chemolithotroph when co- 
cultured with the heterotroph. Application of the 
model assumes that the two types of bacteria can 
coexist in a chemostat receiving two different 
limiting nutrients, one for each microorganism. For 
purposes of illustrating the model, (a) growth of the 
chemolithotroph and the heterotroph are assumed to be 
limited by ferrous iron and an organic acid, 
respectively, and (b) the organic acid is inhibitory 
to the chemolithotroph. In addition, the predictive 
value of the model is dependent on the assumptions 
that there is no product inhibition and the level of 
ferrous iron is not inhibitory to the heterotroph. 
In this model, the steady state biomass of 
T. ferrooxidans is a function of the maximum growth 
rates and substrate half-saturation constants of both 
members of the population as well as the inhibitor 
constant of the chemolithotroph. The model indicates 
that the growth of T. ferrooxidans may be strongly 
linked to the growth" of other organism(s) capable of 
removing organic inhibitors. 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and Reclamation 
and the U. S. Department of the Interior 
(Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), April 
17-22, 1988, Pittsburgh, PA. 
2 

K. L. Shuttleworth is a Graduate 

Assistant and R. F. Unz is Professor of 
Environmental Microbiology, Program in 
Ecology and Department of Civil 



Engineering, The Pennsylvania State 
University, University Park, PA. 



INTRODUCTION 



The microbial aspects of acidic mine 
drainage have been the subject of many 
investigations (Dugan 1975, Harrison 1978, 
Wichlacz and Unz 1981) . One of the key 
reactions in the formation of acidic mine 
drainage is mediated by the 
chemolithotrophic, iron-oxidizing bacterium, 



97 



Thiobacillus ferrooxidans . This organism 
obtains energy from the oxidation of 
ferrous iron or reduced sulfur compounds 
and sustains the oxidation of pyritic 
sulfur by providing a continuous supply of 
the oxidant, Fe (Singer and Stumm 1970). 
Furthermore, hydrolysis of the ferric iron 
yields the mineral acid, H2SO4, which is 
responsible for the low pH of the drainage. 



T. ferrooxidans and other acidophilic 
thiobacilli are of economic importance 
because of their role in the genesis and 
continued production of acidic mine 
drainage; therefore, many researchers have 
studied the effects of inhibitors on the 
thiobacilli (Borichewski 1967, Tuttle and 
Dugan 1976, Tuttle et al . 1977). 
Compounds which have been found to inhibit 
the acidophilic thiobacilli include 
pyruvate, acetate, succinate, glutamate, 
and fumarate. In addition, there has 
been much interest in the inhibitory 
effects of a detergent, sodium lauryl 
sulfate (Dugan 1987, Dugan and Apel 1983, 
Onysko et al . 1984). A proposed scheme 
for the prevention of acidic mine drainage 
involves treating coal refuse with sodium 
lauryl sulfate to inhibit the thiobacilli 
and thus curtail the production of 
pollutants which characterize acidic mine 
waters . 



The focus of most of the acidic mine 
drainage literature has been on the 
acidophilic thiobacilli; however, 
acidophilic heterotrophs , e.g., members of 
the genus Acidiphilium , have recently been 
isolated from acid-mineral environments 
(Harrison 1981, Wichlacz et al . 1986). 
Furthermore, nonacidophilic heterotrophic 
organisms may be found in gelatinous 
streamers in acidic mine drainage (Johnson 
et al. 1979) . 



The organic carbon utilized by 
heterotrophs may include compounds that 
are potentially toxic to the acidophilic 
thiobacilli. It has been shown that 
acidophilic heterotrophs .can use 
succinate, glutamate, citrate, fumarate, 
and other similar carbon sources (Wichlacz 
et al. 1986), some of which have been shown 
also to inhibit T. thiooxidans , T. 
ferrooxidans , or both. Thus a potential 
exists for the two populations to interact 
in the environment through one common 
intermediate . 



MODEL DEVELOPMENT 



A. Background 



Microorganisms may be inhibited by 
various compounds, including microbial 
waste products, high concentrations of 
substrate, and exogenously supplied 
chemicals . Ways in which inhibitors can 
interact with two different snecies are 



shown schematically in figure 1. It is 
possible to develop mathematical models 
which describe these phenomena, and many 
investigators have used such models to 
characterize the effects of product and 
substrate inhibition on microbial growth 
(Davison and Stephanopoulos 1986, Luong 
1987) . Models have been developed which 
describe how product inhibition can be 
relieved by the presence of a second 
organism which removes that product. 
However, there are apparently few models 
which describe possible interactions 
between two species when one of the 
species is affected by an inhibitor which 
is not a product of the system. 



In this paper, a model will be 
developed to express the potential 
interactions between two species of 
acidophilic bacteria, a chemolithotroph and 
a heterotroph, cocultured in a continuous 
culture system. There is no competition 
for a limiting substrate; however, the 
chemolithotrophic organism, T. 
ferrooxidans , is the subject of inhibition 
by an exogenously supplied organic acid 
which the heterotrophic organism, an 
Acidiphilium sp., can degrade. The model 
may predict potential interactions between 
the chemolithotroph and the heterotroph in 
an acidic drainage where exogenous 
inhibitors are present. 



B. Assumptions 



Given that T. ferrooxidans is an 
obligate chemolithotroph and that the 
Acidiphilium sp . is an acidophilic, 
obligate heterotroph, it is possible to 
coculture the organisms in the chemostat 
by using different growth limiting 
nutrients for each organism. The specific 
assumptions are: 

(1) The growth limiting nutrient for T. 

ferrooxidans and for the heterotroph 
will be ferrous iron and an organic 
acid, respectively. 

(2) Growth of T. ferrooxidans will be 
inhibited b~y the organic acid. 

(3) The iron concentration necessary for 
the growth of T. ferrooxidans will not 
be inhibitory to the acidophilic 
heterotroph (Wichlacz 1980). 

(4) There is no product inhibition for 
either organism. 

(5) Given that the mechanism for inhibition 
of T. ferrooxidans by organic acids 
appears to be membrane leakage (Tuttle 
et al 1977) , the model will account for 
an inhibitor which affects both the K, 
and u maK . This is equivalent to the 
noncompetitive inhibition model in 
enzyme kinetics. 

(6) The dilution rate for the chemostat is 
set at a rate which will allow the two 
species to coexist. 



m 



98 



ls ll — l x ll — l p ll 



E^E— ED 



A. Organisms inhibited by 

substrate of other species 



B. Organisms inhibited by high levels of 
their own substrate (substrate 
inhibition) . 



C. Organisms inhibited by product 
of other species 



EMHF ZZ S 



D„ Organisms inhibited by their own 
products (product inhibition) . 



s l 




x l 




p i 

7" 






[5 


\S 


P 2 


r 



E. Organism 2 uses product of organism 1, 
i.e. , P, = S 2 






F. All possible inhibitory effects 



Figure 1 . Schematic of some possible interactions between two species 
with two substrates and two products. Symbols: substrate 1 (Si) ; 
species 1 (X^) ; product 1 (Pi) ; substrate 2 (S 2 ) ; species 2 (X 2 ) ; 
product 2 (P 2 ) ; positive effect ( ); inhibitory effect ( ) 



C. Mathematical symbols 



c (subscript) Chemolithotrophic organism 
(T. ferrooxidans) 



D 



Dilution rate (hr ) 



h (subscript) Heterotrophic organism 
( Acidiphilium sp.) 

I Concentration of inhibitor 

(moles/1) 



K i 



m 



Inhibitor constant 
corresponding to the 
dissociation constant of the 
enzyme-inhibitor complex, 
moles/1 (Michal 1978) 

Half-saturation constant, 
i.e., substrate concentration 
when growth rate is equal to 
% maximum rate (moles/1) 



L Loss term, a function of yield (g/g) 

M Biomass concentration (g/1) 

S Substrate concentration in the 
chemostat (moles/1) 

S Substrate concentration in the feed 
° (moles/1) 

u Specific growth rate (hr ) 

u Maximum specific growth rate (hr ) 
max r & 

D. The Model 



The basis for model development will be 
the Monod model of bacterial growth in a 
chemostat (Tempest 1970) . The Monod model 
is an equation for saturation kinetics which 
describes the growth of microbes in terms of 
a limiting nutrient and factors (K m and 



99 



u ) which are a function of the 
pKy^iology of the organisms. The chemostat 
mimics a simplified natural environment 
with continuous and equal flow rates to and 
from the system. 



In order to interpret data from mixed 
culture work, it is first necessary to 
determine the kinetic parameters (K m and 
Umax) °f the organisms in pure culture. 
Similarily, the K^ for the chemolithotroph 
must be determined from pure culture 
studies on T. ferrooxidans in the presence 
and absence of the exogenous inhibitor. 
When the continuous culture system is at 
steady state, there is no net increase in 
biomass; hence u = D and the relationship 
of specific growth rate to the other 
parameters for the system can be expressed 
for the chemolithotroph and the 
heterotroph in equations (1) and (2) , 
respectively. 



(1) 



u c = D- 



chemolithotroph may be described by 
equation (4) 



(4) 



u c = D 



maxc 



(K ms + S) (1 + I/ Ki ) 



maxc 



- L 



Thus in axenic culture, growth of the 
chemolithotroph will be affected by the 
concentration of inhibitor and the 
inhibitor constant. In mixed culture, the 
concentration of inhibitor will be 
determined in turn by the K . and the 
UjjjQxh of the heterotroph as well as the 
dilution rate. 



The limiting substrate for the 
heterotroph is I and the level of 
limiting substrate at steady state can be 
determined by setting equation (5) equal 
to zero and solving for I. The results of 
this transformation are given in equation 
(6). 



ms 



(2) 



u h = D = 



u. 



maxh 



- I- 



K 



mi 



It is necessary to modify equation (1) 
to include the effects of a noncompetitive 
inhibitor. In a Lineweaver- Burke 
linearization of saturation kinetics, it is 
easy to distinguish the effects of an 
inhibitor by the apparent K m and u max . 
When these inhibitor effects are accounted 
for and the equation is rearranged into 
standard form, the result is equation (3) . 



(3) 



max 



(K m + S) (1 + I/K t ) 



This equation is easily derived and 
has been credited to Haldane who apparently 
first described it in 1930 (Luong 1987) and 
to Aiba and Shoda (Davison and 
Stephanopoulos 1986) . The accuracy of the 
completed model will be highly dependent on 
the ability of equation (3) to predict the 
effects of the inhibitor. It should be 
noted that Luong (1987) believes equation 
(3) fails to adequately predict inhibitor 
effects when the inhibitor and the 
substrate are the same compound. However, 
this equation served as the basis for the 
model of Davison and Stephanopoulos (1986) 
for product inhibition. Furthermore, the 
inhibitor of T. ferrooxidans is not an 
available substrate for the organism. 



(5) 



(6) 



dM u . 
— = maxh 

dt 



I * M 



K . + I 
mi 



1 = K mi < D + L h> 



L h M 



DM,. 



u 



maxh 



- (D + L. ) 



Substituting equation (6) into (4) 
yields equation (7) which shows the 
relationship between the chemolithotroph 
and the heterotroph. 



As stated above, the model given in 
equation (7) includes the simplifying 
assumption that (I) is not inhibitory to 
the heterotroph at any concentration. 
However, acidophilic heterotrophs can be 
inhibited by high levels of organics 
(Harrison 1981). Luong (1987) has 
suggested that equation (8) represents a 
good model for substrate inhibition. S 
is the maximum substrate level which does 
not cause inhibition and n is determined 
empirically. In order to use this model, 
one would have to solve equation (8) for S 
and substitute this back into equation (4) 
as given above. However, solving equation 
(8) for S by setting dM/dt = is 
complicated and beyond the scope of this 
paper. Hence, the model developed here is 
appropriate only when I <S m . This appears 
to be reasonable since tR~e levels of 
organics in acidic mine drainage are 
typically very low (0.0037c,; Johnson et al. 
1979) and the high levels (>0.57o) of 
organics referred to by Harrison (1981) are 
typical for culture media used in 
conventional microbiology. 



Equation (3) must be corrected to 
include loss of cellular material due to 
cell maintenance. Again steady state will 
be assumed; therefore, the behavior of an 
axenic culture of the inhibited 



It is of interest to learn how the 
inhibitor affects the steady state 
population of the chemolithotroph. This 
requires that equation (9) be set equal to 



100 



(7) u =/ 
c 



(K + S 

ms 



u * S 

maxc 



L K i (maxh " < 



V L - D 
c 



(D + L h )> 



(8) dM u * S 

= max (1 _ s/s .n 

M * dt K + S u s/ m ; 

m 



(9) dM u * M * S 

— - = maxc - (D + L )M 

dt (K ms + S) (1 + I/K ± ) 

(10) S = K ms< D + L c> 0; + I / K i ) 

u - (D + L ) (1 + I/K.) 
maxc x cr i' 



(11) M = D < S o - S > 



(D + L c ) 



D * S 

(12) M = ^o 

c 



D ,,. K |(D + L ) (1 + I/K.)l 



(D + L ) (D + L ) 



maxc 



[(D + L c ) (1 + I/KJ 



n * q 

(13) M = _ ^o 

c 



(D + L c ) (D + L c ) 



(X) where X equals 



ms 



[(D + L c ) 



(*1 fcnaxh - < D + \1 4 
|(D + L c ) (K, [u maxh - (D + L h )] 



K mi (D + Lj 



^J 



maxc 



+ K . (D + 
mi v 



zero. When equation (9) is solved for S, 
the result is equation (10) . The steady 
state biomass is a function of the dilution 
rate, maintenance loss, initial substrate 
level, and the steady state substrate level 
and can be expressed by equation (11) . 
Substituting equation (10) into equation 
(11) gives equation (12) which shows the 
relationship between the inhibitor level , 
the inhibitor constant, and the steady 
state biomass. However, (I) is again 
determined by the population of the 
heterotroph as given in equation (6) . 
Therefore, equation (6) should be 
substituted for the value of (I) in 
equation (12) . This results in equation 
(13) which could be simplified by combining 
terms and renaming them by definition. 
However, in its present form one can see 
that the steady state biomass of the 
chemolithotroph depends on the following 
physiological characteristics of the two 
microbes; the maximum growth rate of both 
the chemolithotroph and the heterotroph, 
the half saturation constants of both the 
chemolithotroph and the heterotroph, and 
the inhibitor constant of the 
chemolithotroph. In contrast, the steady 
state biomass of the heterotroph would 
depend only on the Uuj ax and K m for the 
heterotroph itself. Biomass of both the 
heterotroph and the chemolithotroph would, 
however, also depend on the yields and 



dilution rate, 



SUMMARY AND CONCLUSIONS 



A mathematical model was developed to 
express the potential relationship between 
the chemolithotroph, T. f errooxidans , and 
a heterotroph, Acidiphilium sp. , in the 
presence of a compound which inhibits T. 
ferrooxidans but is the growth substrate 
for the heterotroph. In the model, the 
growth of the chemolithotroph was strongly 
linked to the growth of the heterotroph. 
The model must now be verified by actual 
trials with the chemostat. If the model 
does not fit the data, the model must be 
altered. The key assumptions which may 
need to be changed are: (1) there is no 
substrate inhibition of the heterotroph, 
(2) the mechanism of inhibition of the 
chemolithotroph is noncompetitive, (3) 
there is no product inhibition, and (4) 
no other interactions exist between the 
two species. 



Although the extent of inhibitory 
effects of organics on T. ferrooxidans in' 
the natural environment is not known, 
Dispirito et al. (1981) have reported that 
a soluble component of pyrite was capable 



101 



of inhibiting the growth of T. ferro- 
oxidans. Furthermore, it has already been 
established that obligate heterotrophs or 
mixotrophic thiobacilli capable of growing 
in an acidic environment can reduce the 
inhibitory effects of organics on T. 
ferrooxidans (Harrison 1984, Nerkar et al . 
1977, Unz and Wichlacz 1982). 

This model should also be generally 
applicable to situations where one 
organism metabolizes or transforms a 
substance which is inhibitory to a second 
organism. There are many such cases in the 
microbial world, including the ability of 
Beggiatoa to detoxify (oxidize) HoS 
(Atlas and Bartha 1987) and the ability of 
chlorophenol-degrading bacteria to 
detoxify the environment for other 
microbes and for higher organisms (Steiert 
and Crawford 1985) . 

LITERATURE CITED 

Atlas, R. M. and R. Bartha. 1987. 

Interactions among microbial pop- 
ulations, p. 101-132. In R. M. Atlas 
and R. Bartha, MicrobiaT Ecology: 
Fundamentals and Applications, second 
edition. Benjamin/ Cummings Publ. Co. 
Inc., Reading, MA. 

Borichewski, R. M. 1967. Keto acids as 

growth-limiting factors in autotrophic 
growth of Thiobacillus thiooxidans . 
J. Bacterid . 93:597-599^ 

Davison, B. H. and G. Stephanopoulos . 1986. 
Coexistence of S_. cerevisiae and E. 
coli in chemostat under substrate 
competition and product inhibition. 
Biotechnol. Bioeng. 28:1742-1752. 

Dispirito, A. A., P. R. Dugan , and 0. H. 

Tuovinen. 1981. Inhibitory effects of 
particulate materials in growing 
cultures of Thiobacillus ferrooxidans . 
Biotechnol. Bioeng. 23:2761-2769. 

Dugan, P. R. 1975. Bacterial ecology of 

strip mine areas and its relationship 
to the production of acidic mine 
drainage. Ohio J. Sci. 75:266-279. 

Dugan, P. R. 1987. Prevention of formation 
of acid drainage from high sulfur coal 
refuse by inhibition of iron- and 
sulfur- oxidizing microorganisms. II. 
Inhibition in "Run of Mine" refuse 
under simulated field conditions. 
Biotechnol. Bioeng. 29:49-54. 

Dugan, P. R. and W. A. Apel. 1983. Bacteria 
and acidic drainage from coal refuse: 
inhibition by sodium lauryl sulfate 
and sodium benzoate. Appl. Environ. 
Microbiol. 46:279-282. 

Harrison, A. P. Jr. 1978. Microbial 

succession and mineral leaching in an 
artificial coal spoil. Appl. Environ. 
Microbiol. 36:861-869. 



Harrison, A. P. Jr. 1981. Acidiphilium 
cryptum gen. nov. sp. nov. , 
heterotrophic bacterium from acidic 
mineral environments. Int. J. Syst. 
Bacteriol. 31:327-332. 

Harrison, A. P. Jr. 1984. The acidophilic 
thiobacilli and other acidophilic 
bacteria that share their habitat. 
Ann. Rev. Microbiol. 38:265-292. 

Johnson, D. B., W. I. Kelso, and D. A. 
Jenkins. 1979. Bacterial streamer 
growth in a disused pyrite mine. 
Environ. Pollut. 18:107-118. 

Luong, J. H. T. 1987. Generalization of 

Monod kinetics for analysis of growth 
data with substrate inhibition. 
Biotechnol. Bioeng. 29:242-248. 

Michal, G. 1978. Determination of 

Michaelis constants and inhibitor 
constants, p. 29-40. In H. U. 
Bergmeyer (ed.), Principles of Enzy- 
matic Analysis. Verlag Chemie, New 
York. 

Nerkar, D. P., U. S. Kumta, and N. F. 

Lewis. 197 7. Pyruvate inhibition of 
Ferrobacillus ferrooxidans reversed 
by Rhodotorula yeast. J. Appl. 
Bacteriol. 43:117-121. 

Onysko, S. J., R. P. Kleinmann, and P. M. 
Erickson. 1984. Ferrous iron oxida- 
tion by Thiobacillus ferrooxidans : 
Inhibition with benzoic acid, sorbic 
acid, and sodium lauryl sulfate. Appl 
Environ. Microbiol. 48:229-231. 

Singer, P. C. and W. Stumm. 1970. Acidic 
mine drainage: The rate limiting 
step. Science 167:1121-1124. 

Steiert, J. G. and R. L. Crawford. 1985. 
Microbial degradation of chlorinated 
phenols. Trends in Biotechnol. 
3:300-305. 

Tempest, D. W. 1970. The continuous 

cultivation of micro-organisms. I. 
Theory of the chemostat, p. 259-276. 
In J. R. Norris and D. W. Ribbons 
Teds . ) , Methods in Microbiology, 
Vol. 2. Academic Press, New York. 

Tuttle, J. H. and P. R. Dugan. 1976. 

Inhibition of growth, iron, and sulfur 
oxidation in Thiobacillus ferrooxidans 



by simple organic compounds, 
Microbiol. 22:719-730. 



Can. J. 



Tuttle, J. H. , P. R. Dugan, and W. A. Apel, 
1977. Leakage of cellular material 
from Thiobacillus ferrooxidans in 
the presence of organic acids. 
Appl. Environ. Microbiol. 3_3: 459-469 , 



102 



Unz, R. F. and P. L. Wichlacz. 1982. 

Microbiology of coal mine drainage 
treatment, p. 199-208. In A. M. Al 
Taweel (ed.), Proc. 64th CIC Coal 
Symp., Canada Inst. Conf . , Ottawa. 

Wichlacz, P. L. 1980. Acidophilic hetero- 
trophic bacteria from acid mine 
drainages in central Pennsylvania. 
M. S. thesis. The Pennsylvania State 
University, University Park, PA. 

Wichlacz, P. L. and R. F. Unz. 1981. 

Acidophilic, heterotrophic bacteria 
of acidic mine waters. Appl. Environ, 
Microbiol. 41:1254-1261. 

Wichlacz, P. W. , R. F. Unz, and T. A. 
Langworthy. 1986. Acidiphilium 
angustum sp. nov. , Acidipnilium 
facilis~ sp . nov., and Acidiphilium 
rubrum sp . nov.: acidophilic hetero- 
trophic bacteria isolated from acidic 
coal mine drainage. Int. J. Syst. 
Bacterid. 36:197-201. 



103 



REHABILITATION OF WASTE ROCK DUMPS AT THE RUM JUNGLE MINE SITE 1 



John W. Bennett, John R. Harries and A. Ian M. Ritchie 2 



Abstract. —The release of acid, copper, manganese and zinc was a 
major environmental problem at the abandoned Rum Jungle mine site in 
the Northern Territory, Australia. The main sources of pollution were the 
three waste rock dumps and the heap leach pile, all containing pyritic 
material. The site was rehabilitated between 1982 and 1986 and, as part of 
this program, the dumps were reshaped and covered with a three-layer 
cover, including a compacted clay layer. A monitoring program has been 
carried out to assess the effectiveness of the works on two of the dumps. 
Lysimeters installed in the dumps have shown that less than 5 % of the 
incident rain percolates through the covers. The distribution of heat 
production in the dumps has been derived from measured temperature 
distributions and shows that the rate of pyritic oxidation was greatly 
reduced by emplacement of the covers. Comparison of oxygen concentra- 
tions in the two dumps before and after rehabilitation shows that the covers 
have greatly reduced the transport of oxygen into the dumps and effectively 
stopped thermal convection. 



INTRODUCTION 

There are many surface mine sites where acid drainage 
and the release of heavy metals create a major environmental 
problem which can continue long after mine operations cease. 
The problems of acid drainage occur at many mine sites around 
the world. There is an urgent need to know the effectiveness of 
techniques for controlling the release of these pollutants. At 
present there is inadequate knowledge and experience to allow 
the advantages, disadvantages and costs of various rehabilita- 
tion techniques to be compared. 

Opencut mining to extract uranium/copper ore was 
carried out at the Rum Jungle mine in the Northern Territory, 
Australia, between 1954 and 1964, and the site was abandoned 
in 1971. However acid, copper, manganese and zinc continued 
to be released into the local river system (Northern Territory 
Department of Mines and Energy, NTDME 1987). The climate 



'Paper presented at the 1988 Mine Drainage and Surface Mine 
Reclamation Conference sponsored by the American Society for 
Surface Mining and Reclamation and the U. S. Department of 
the Interior (Bureau of Mines and Office of Surface Mining, 
Reclamation and Enforcement), Pittsburgh, PA, April 17-22, 
1988. 

2 John W. Bennett is a Research Scientist, John R. Harries is a 
Principal Research Scientist, A. Ian M. Ritchie is Leader, 
Physics of Environment Section, Environmental Science Divi- 
sion, Australian Nuclear Science and Technology Organisation 
(ANSTO), PMB 1, Menai, NSW 2234, Australia. Prior to April 
1987 ANSTO was the Australian Atomic Energy Commission. 



at the site is tropical with most of the 1.5 m annual rainfall 
occurring in the wet season from December to March. 

The main sources of pollution were the three waste rock 
dumps and an abandoned heap leach pile, all containing pyritic 
material which oxidised in the presence of oxygen and water to 
produce sulfuric acid and heavy metal salts. These pollutants 
were leached by rainwater percolating down through the dumps 
and carried by groundwater to the river. Other sources of 
pollution were three opencuts, two filled with polluted water 
and one largely filled with tailings. The site was rehabilitated 
between 1982 and 1986. This paper discusses the results from 
monitoring two of the dumps (White's and Intermediate) and 
compares conditions before and after rehabilitation. Gibson and 
Pantelis (1988 , elsewhere in these proceedings) describe the 
groundwater monitoring being carried out at the site. 

Probe holes were drilled down to the original ground 
surface in White's and Intermediate dumps (fig. 1) to provide 
access for measuring temperature, gas composition and water 
content. Three types of liner have been used. Type "n" holes 
(six on White's and two on Intermediate dumps) are lined with 
a 50 mm i.d. polyethylene pipe and sealed at the bottom to 
exclude water. Type "g" holes have gas ports inset into the liner 
at depths of 0.5, 1.0, 1.5, 2.0, 3.0, 5.0 m and then at 2.5 m 
intervals to the bottom of the hole. Type "p" holes have pairs of 
gas sampling tubes attached to the outside of the polyethylene 
liner to directly sample the gas in gravel-filled zones at each 
meter down the hole. 

The first drilling program was in 1976 when six n-holes 
were installed in White's dump. In 1982 three g-holes were 



104 



EAST BRANCH FINNISS 




100m 



100m 



White's dump before rehabilitation. 



White's dump after rehabilitation. 





Intermediate dump before rehabilitation. 



Intermediate dump after rehabilitation. 



Figure 1. -White's and Intermediate dumps showing positions of probe holes before and after rehabilitation. 
The n-holes are shown as circles, the g-holes as diamonds, and the p-holes as triangles. 



installed in White's dump. The hole in White's dumps were 
preserved during the earthworks which took place on White's 
between August 1983 and June 1984, although the length of the 
liners of most holes had to be adjusted to conform with the 
recontoured dump. Finally ten p-holes were installed in White's 
dump in June 1987 (fig. 1). 

The first drilling program on Intermediate dump was in 
1982 when four g- and two n-holes were installed. In 1984, nine 
p-holes were installed in Intermediate dump to provide more 
information on temperature and pore gas composition profiles 
before rehabilitation. All the holes in Intermediate dump were 
lost when Intermediate was rehabilitated between August and 
December 1985. Nineteen new p-holes were installed in Inter- 
mediate dump in November and December 1985 after the 
earthworks on the dump were completed (fig. 1). 



REHABILITATION 

The Rum Jungle site was rehabilitated between 1983 and 
1986 to reduce the level of pollutants in the river, reduce public 
health hazards, reduce pollution levels in the flooded open cuts, 
and achieve aesthetic improvements including revegetation 
(NTDME 1987). The total cost was $(Aust)18.6 million of 
which $2.8 million was spent rehabilitating the three waste rock 
dumps (in June 1986 $(Aust)1.00 = $(US)0.68). The areas of 
White's, Intermediate and Dyson's dumps before rehabilitation 
were 26, 6.9 and 8.4 ha respectively. 

The treatment applied to the waste rock dumps was 
designed to reduce the ingTess of water and the release of 
pollutants. First the dumps were reshaped so that the tops had a 
maximum slope of 5° and and the sides a maximum slope of 1 



105 



in 3 horizontal. The tops were then covered with a three-layer 
cover consisting of a layer of compacted clay (minimum 
thickness 225 mm) on the bottom as a moisture barrier, a layer 
of sandy clay loam (minimum thickness 250 mm) as a moisture 
retention zone to support vegetation and prevent the clay layer 
drying out, and a layer of gravelly sand (minimum thickness 
150 mm) on the top to provide erosion protection and act as a 
pore breaking zone to restrict moisture loss by evaporation in 
the dry season. A similar three-layer cover was emplaced on the 
sides of the dumps, but the layers were thicker (minimum 
thickness 300 mm of compacted clay and minimum thickness 
300 mm of sandy clay loam) and crushed rock was used for the 
erosion barrier (minimum thickness 150 mm). Engineered 
runoff channels and erosion control banks were constructed on 
the tops and sides of the dumps. Vegetation was established to 
stabilise the dump surface against the long-term effects of 
erosion. 

HEAT PRODUCTION 

The oxidation of pyrite to sulfuric acid and ferrous sulfate 
is exothermic and releases 1440 kJ mol-'FeS 2 . It is unlikely that 
there are other reactions occurring which release significant 
amounts of heat. The released heat causes elevated temperatures 
in the regions where pyritic oxidation is occurring. Before 
rehabilitation, the temperatures at some locations within the 
dumps exceeded 50°C. A one-dimensional heat transfer model 
has been used to derive the distribution of heat production in the 
dumps from vertical temperature profiles measured before and 



after rehabilitation (Harries and Ritchie 1980, 1987). The rate of 
oxidation of pyrite can be obtained directly from the heat source 
distribution using the heat of reaction. 

Before rehabilitation, heat production was occurring in 
White's dump at depth in holes A, C, D and F (fig. 2). After 
rehabilitation, the heat production was either very low or zero in 
all holes, with the possible exception of hole C. Comparison of 
heat production distributions before and after rehabilitation 
shows that the oxidation occurring before rehabilitation was 
effectively stopped by rehabilitation. 

PORE GAS COMPOSITION 

Before rehabilitation, the supply of oxygen was the main 
process limiting the rate of oxidation in the Rum Jungle waste 
rock dumps. The oxygen profiles showed that oxygen was 
transported to the oxidation sites by a combination of diffusion, 
thermal convection, and advection driven by variations in 
atmospheric pressure (Harries and Ritchie 1985). Each of these 
oxygen transport processes leads to a characteristic oxygen 
concentration profile. Thermal convection causes the oxygen 
concentration to be higher near the base of the dump. Diffusion 
causes the oxygen concentration to decrease monotonically with 
depth. Advection driven by variations in the atmospheric 
pressure leads to short-term changes in the oxygen concentra- 
tion over timescales of less than a day. Increasing pressure 
causes air to flow into the pore space and, because the incoming 
air has a higher oxygen content than air already in the dump, the 




1.6 



1.2 



0.8 



0.4 



(b) 



8.0 12.0 

Depth (m) 



20.0 




8.0 12.0 

Depth(m) 



20.0 



Figure 2. -Heat sources profiles derived from temperature profiles measured in White's dump (a) before 
rehabilitation (Mar 1979 - Sep 1983), (b) after rehabilitation (Dec 1984 - Sept 1986). 



106 



oxygen concentration measured at a given point increases. At 
tropical locations like Rum Jungle, the dominant atmospheric 
pressure variations are atmospheric tides which have two 
maxima and two minima a day. This causes the oxygen 
concentration at a given point in the dump to have two maxima 
and two minima a day. 

Carbon dioxide concentrations measured in the dumps 
tended to be anticorrelated with the oxygen concentrations. This 
suggests that the controlling process was the rate of exchange 
between the interstitial gas and the atmosphere; the lower the 
exchange rate between the interstitial gas and the atmosphere 
the greater the time for the oxygen to be used in the oxidation 
process and the smaller the opportunity for carbon dioxide to 
escape into the atmosphere. 

The air permeability in the rehabilitated Intermediate 
dump is high near the surface, lower at mid-depth, and increases 
at some locations near the base. 

The oxygen concentration in the pore gas decreased soon 
after the installation of the compacted clay cover. Figures 3 and 
4 show the distribution of oxygen in Intermediate and White's 
dumps before and after rehabilitation. The tongue of oxygenated 
air evident at depth in both dumps before rehabilitation indi- 
cates that thermal convection was transporting oxygen from the 
sides of the dumps and up through the hot regions. Since 
rehabilitation, the oxygen concentrations have been low at 
depth at all measuring points except in the northwest corner of 
White's dump where there are very low levels of pyrite. The 
clay cover effectively stopped oxygen transport by thermal 
convection, and greatly reduced diffusion and atmospheric 
pressure-driven advection. 

However, after the covers had been in place for about a 
year, the oxygen concentrations in the top few meters were 
found to increase in the morning and evenings in the wet season 
(fig. 5). This behavior is characteristic of advection driven by 
variations in atmospheric pressure. The fact that these elevated 
oxygen concentrations were only present in the wet season can 
be explained by considering the effect of seasonal changes in 
the permeability of the compacted clay cover on the transport of 
air into the dump both near and away from the holes. If the clay 
near the holes was not as well compacted as the clay further 
from the holes, airflow caused by variations in the atmospheric 
pressure would tend to be concentrated in the higher permeabili- 
ty material near the holes. 



Zg K Yg 



OCT 1984 



80 m 
AHD 



70 m 




J 



100 m 



200m 300m 

SEPT 1986 




Figure 3. —Oxygen distributions in Intermediate dump (a) Oct 
1984 before and (b) Sept 1986 after rehabilitation. 



100m.- 
AHD 



80m 




J I L 



SEPT 1987 




Figure 4. -Oxygen distributions in White's dump (a) Apr 1983 
before and (b) Sept 1987 after rehabilitation. 



The low oxygen concentrations and the lack of diurnal 
variation in the dry season indicate that the clay near the holes 
has a similar permeability to that of the clay further away. This 
suggests that there is cracking of the clay layer in the dry 
season, and that the cracks provide paths over the whole surface 
for advection of air by atmospheric pressure variations. The 
reappearance of the diurnal variations in the oxygen concentra- 
tions early in the wet season shows that most of the cracks close 



^20- 



o 
> 



> 
o 



10 



"i 1 1 1 1 1 r 

8 March 1987 



Oxygen 




cc 

I/) 
I/) 

LU 

a. 



998 



1200 1500 1800 

Atmos Pressure 




0600 



0900 



1200 
TIME OF DAY 



1500 



1800 



Figure 5. —Diurnal changes in the oxygen concentrations at 

various depths in Hole 2 in Intermediate dump during the 
wet season. 



107 



as the moisture content of the clay increases, but the clay near 
the holes does not seal as well as that further away. 

Oxygen concentrations at depth in the dumps continue to 
be much less than they were before rehabilitation. The compact- 
ed clay cover does appear to have stopped oxygen transport by 
thermal convection. 

WATER BALANCE 

The main aim of covering the dumps was to reduce 
ingress of water and thereby reduce the release of pollutants. 
Sets of lysimeters were installed in the reshaped White's and 
Intermediate dumps before emplacement of the clay layer. The 
lysimeters consist of 200 L drums with pipes to allow any water 
collected to be extracted and measured. The drums are filled to 
300 mm with gravel to make a water collection zone and the 
rest of the space is filled with dump material. Theoretical 
models of unsaturated waterflow have been used to verify the 
effectiveness of the lysimeters in collecting infiltrating water 
(Gibson 1987). 

The amount of water collected by the ten lysimeters in 
White's dump in each of the three wet seasons between 1985 
and 1987 is equivalent to between 2.0 and 2.5% of the incident 
rain. Lysimeters in Intermediate dump have only been opera- 
tional for one wet season and the amount of water collected 
corresponds to 4.8 % of the incident rain. Before rehabilitation 
it was estimated that about 50% of the incident rain percolated 
through the dumps. These results indicate that the compacted 
clay cover is achieving the desired reduction in water ingress. 

CONCLUSIONS 

Monitoring the waste rock dumps at Rum Jungle has 
shown that rehabilitation by reshaping and covering with 
compacted clay has been effective in greatly reducing the 
ingress of water, the rate of oxidation of pyrite and the transport 
of oxygen. Further monitoring will be necessary to check that 
the reduction in pollution generation and water infiltration rates 
continue in the longer term. 

The monitoring program on the waste rock dumps has 
allowed a quantitative assessment to be made of the effective- 
ness of the techniques used to reduce pollution loads from 
dumps. A more complete quantitative assessment will be 
possible when the effects of rehabilitation on ground and 
surface waters have been determined. However, it was recog- 
nised that improvements in the quality of ground and surface 
water at the mine site could take some years (Pantelis 1987, 
Gibson and Pantelis 1988). The lysimeters and gas composition 



measurements on White's dump in particular were seen as a 
means of acquiring early evidence of any gross shortcomings in 
the rehabilitation strategy which would allow the strategy to be 
applied to the other dumps to be modified if necessary. The 
reduced ingress of water and the low or zero oxidation rate 
(pollution generation rate) gives confidence that the release of 
pollutants from the waste rock dumps has decreased. 

Monitoring was considered to be an important part of the 
Rum Jungle rehabilitation project and it was costed as part of 
the project. As well as the program on the waste rock dumps 
described in this paper, there are monitoring programs on 
groundwater, surface waters and vegetation. The results from all 
these programs will allow the cost-effectiveness of the Rum 
Jungle rehabilitation project to be assessed and used as a 
benchmark for other projects to rehabilitate mine sites where 
there are pyritic wastes. 

LITERATURE CITED 

Gibson, D. K. 1987. Mathematical study of a lysimeter. 

ANSTO/E666. 10 p. Australian Nuclear Science and 
Technology Organisation, Menai, Australia. 

Gibson, D. K. and Pantelis G. 1988. Forecasting the effect of 
mine site rehabilitation works on local groundwater 
quality. In: Proceedings 1988 Mine Drainage and Surface 
Mine Reclamation Conference, Amer. Soc. Surface Min. 
and Reclam. and U.S. Dept of Interior. 

Harries, J. R. and Ritchie A. I. M. 1980. The use of temperature 
profiles to estimate the pyritic oxidation rate in a waste 
rock dump from an opencut mine. Water Air Soil Pollut. 
15:405-423. 

Harries, J. R. and Ritchie A. I. M. 1985. Pore gas composition 
in waste rock dumps undergoing pyritic oxidation. Soil 
Sci. 140:143-152. 

Harries, J. R. and Ritchie A. I. M. 1987. The effect of 

rehabilitation on the rate of oxidation of pyrite in a mine 
waste rock dump. Environ. Geochem. Health 9:27-36. 

NTDME 1987. The Rum Jungle Rehabilitation Project, Final 
Project Report. 346 p. Northern Territory Department of 
Mines and Energy, Darwin, Australia, June 1987. 

Pantelis, G. 1987. Modelling water and contaminant transport in 
the Rum Jungle mine overburden heaps. AAEC/E653. 17 
p. Australian Atomic Energy Commission, Menai, Aus- 
tralia. 



108 



CHEMICAL INHIBITION OF IRON-OXIDIZING BACTERIA IN WASTE ROCK 
AND SULFIDE TAILINGS AND EFFECT ON WATER QUALITY 1 



George R. Watzlaf 2 



Abstract. — The effectiveness of sodium lauryl sulfate 
(SLS), potassium benzoate, and potassium sorbate in controlling 
the population of iron-oxidizing bacteria, thereby reducing acid 
production, was tested on sulfide tailings and waste rock, 
common waste products of metal mining. The waste rock was 
unweathered and contained 4.07 percent total sulfur, all in the 
pyritic form. Two different samples of sulfide tailings were 
used, an extensively weathered material (5.98 percent total 
sulfur, 0.17 percent pyritic sulfur, and 5.81 percent sulfate 
sulfur) and a slightly weathered material (20.57 percent total 
sulfur, 19.73 percent pyritic sulfur, and 0.84 percent sulfate 
sulfur). Two sample sizes were used in the experiment, 7 kg and 
100 kg. The 100-kg samples of each material were treated once 
with a 600-mg/kg dose of SLS or potassium benzoate. The 7-kg 
samples were treated with a high (600 mg/kg) or low (60 mg/kg) 
dose of SLS, potassium benzoate, or potassium sorbate. The 
treated samples and untreated control samples were subsequently 
leached once per week with filtered demineralized water (an 
amount equivalent to 2.5 cm of precipitation). For the 100-kg 
samples of the waste rock, a single treatment of SLS and 
potassium benzoate completely inhibited iron-oxidizing bacteria 
repopulation for 182 and 231 days, respectively. Acidity in the 
leachate from the 100-kg samples treated with SLS and benzoate 
remained below the untreated control for 287 and 343 days, 
respectively. Similar results were obtained from the 7-kg 
samples of waste rock. In the extensively weathered sulfide 
tailings, none of the treatments inhibited the iron-oxidizing 
bacteria or reduced acidity levels in the leachate. In the 
slightly weathered sulfide tailings, all treatments reduced the 
bacterial populations, but did not significantly reduce acid 
production. However, in additional tests of the slightly 
weathered tailings, the removal of the weathered products prior 
to treatment with SLS or benzoate resulted in lower populations 
of iron-oxidizing bacteria and reduced acidity levels. 



1 Paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

2 George R. Watzlaf is Environmental Engineer, 
U.S. Bureau of Mines, Pittsburgh Research Center, 
Pittsburgh, PA. 



INTRODUCTION 

The U.S. Bureau of Mines is currently 
researching at-source control of acid mine drainage 
(AMD). AMD is formed by water and oxygen reacting 
with the sulfide minerals that are associated with 
coal and metal mining wastes. A common remedy to 
AMD i3 chemical treatment of the contaminated water 
with alkaline materials such as lime, limestone, 
sodium hydroxide, or sodium carbonate. These 
treatments raise pH, reduce acidity, and 
precipitate metals. However, chemical 



109 



neutralization is expensive, may be necessary for 
many years, and generates large volumes of sludge. 

A more direct approach is to stop or slow 
sulfide mineral oxidation and consequent water 
contamination. The chemical reactions involved in 
the formation of AMD can be summarized as follows: 

2FeS 2 + 70 2 + 2H 2 «e* 2Fe 2+ + 4SOi, 2 " + 4H + , (1) 

lFe 2 + + 2 + 4H + V" lFe3 + + 2H 2 0, (2) 

FeS 2 + 1 HFe3 + + 8H 2 •*»- 15Fe 2+ + 2S0n 2_ + 16H + . (3) 

The sulfide moiety of pyrite (FeS 2 ) can be oxidized 
by oxygen (Eq. 1) resulting in ferrous iron (Fe 2+ ), 
sulfate (SO4 2- ), and acid (H + ) . Ferrous iron can 
then be oxidized to ferric iron (Fe3 + ), which can 
directly oxidize pyrite (Eq. 3). This results in 
additional ferrous iron, sulfate, and acid. The 
rate-limiting step of FeS 2 oxidation is equation 2, 
which proceeds slowly at low pH in the absence of 
iron-oxidizing bacteria such as Thiobacillus 
ferrooxidans (Singer and Stumm 1970). These 
bacteria, which are indigenous to areas where 
pyrite occurs, can accelerate the rate of equation 
2 as much as 1 x 10° times (Singer and Stumm 1970). 
The catalytic activity of these bacteria may be the 
cause of over 80 pet of the AMD problem in the 
United States (Browning 1970). 

Various chemicals can be used to inhibit 
T. ferrooxidans . Sodium lauryl sulfate (SLS) and 
other surfactants have been shown to reduce AMD in 
laboratory-scale, pilot-scale and full-scale field 
tests on coal refuse (cf. Stancel 1982, Kleinmann 
and Erickson 1982). A disadvantage of applying 
surfactants in solution form is the necessity to 
intermittently repeat the application. However, 
even with their short effective duration and need 
to reapply every three to six months, anionic 
surfactants can be more cost effective than 
treatment of the original highly acidic water with 
conventional alkaline materials. To remedy the 
need to reapply the surfactant, Kleinmann et al . 
(1980) developed a method using surfactant- 
impregnated rubber pellets that slowly release a 
surfactant over a longer period of time. 

In an attempt to identify chemicals that have 
a longer effect when added as solutions, organic 
acids and their ability to form sparingly-soluble 
salts were studied (Onysko et al . 1981a). It was 
hypothesized that these salts would precipitate on 
pyrite surfaces and that these precipitates would 
dissolve upon decrease in pH to the biologically 
active form. This scheme was intended to extend 
the inhibitory effects of these chemicals. Salts 
of benzoic acid and sorbic acid were deemed 
appropriate for study (Onysko et al . 1984b). In 
laboratory and pilot-scale tests on coal refuse, 
benzoate and sorbate were found to be effective 
bacterial inhibitors; however, no increase in 
duration of effective inhibition was noted relative 
to conventional SLS solution application (ibid). 

Potassium benzoate and potassium sorbate at 
low concentrations should have no adverse effects 
on the aquatic environment. These chemicals 
specifically inhibit acidophilic microorganisms 
(Eklund 1983, Eklund 1980, Cruess and Richert 1929, 
and Sofos and Busta 1981) and are on the Food and 
Drug Administration Generally Regarded as Safe 
(GRAS) List as suitable for inclusion in foods and 
beverages (U.S. Code of Federal Regulations 1983). 
On the other hand, small concentrations of SLS may 



be deleterious to fish and other aquatic life 
(Margaritis and Creese 1979). Although this 
potential problem exists with SLS, it has never 
been detected in any waterway near any of the 
refuse sites at which the Bureau of Mines has 
supervised SLS application (Kleinmann and Erickson 
1983). 

Because various metal sulfides undergo similar 
chemical reactions (cf. Onysko 1985), inhibition of 
iron-oxidizing bacteria should also reduce acid 
production from these wastes. Therefore, the 
objectives of this study were to determine if these 
chemicals (surfactants and organic acids) would 
inhibit the iron-oxidizing bacteria, and if 
inhibition of these bacteria would reduce acid 
production. 



WASTE MATERIALS 



The sulfide taili 
came from a metal mine 
from this mine is pass 
processed for recovery 
gold, and silver. The 
stages of crushing, gr 
waste product from the 
the sulfide tailings ( 
The extensively weathe 
experiment were taken 
tailings pile. The si 
were taken from under 
pile. 



ngs used in this experiment 

in Quebec, Canada. The ore 
ed through a concentrator and 
of copper, zinc, pyrite, 
ore passes through multiple 
inding, and flotation. The 

flotation cells constitutes 
70 pet less than 200 mesh), 
red tailings used in the 
from the surface of a large 
ightly weathered tailings 
the oxidized layer of the 



The waste rock came from a mine in British 
Columbia, Canada, where the sulfide ore is mined 
for its silver, copper, and gold content. The 
material used in the experiment was unprocessed 
because it was below ore grade. The material was 
unsized (15 cm x cm) and represented the most 
acid-producing material from this mine. 

Samples of each material were taken at the 
beginning of the experiment. Total sulfur was 
measured using the combustion furnace method 
(American Society for Testing and Materials 1983). 
To differentiate the sulfur forms, two acid 
extractions (hydrochloric acid and nitric acid) 
were performed. The residues from these extrac- 
tions were analyzed in a combustion furnace, and 
the percentages of total, pyritic, sulfate, and 
organic sulfur were calculated (table 1). 



EXPERIMENTAL PROCEDURES 

The supply of waste materials limited the 
experiment to three 1 00-kg samples of each 
material: one untreated control, one treated with 
SLS, and one treated with potassium benzoate. In 
addition to these 1 00-kg samples, eight 7-kg 
samples of each material were treated as follows: 
two untreated controls, a high and low dose of SLS, 
a high and low dose of potassium benzoate, and a 
high and low dose of potassium sorbate. The 
experimental design is summarized in table 2. 

For the 100-kg samples, nine 200-L plastic 
barrels were set up on wooden platforms. Two holes 
were drilled in the bottom of each barrel, and a 
plastic bucket was placed underneath each hole to 
collect leachate. For the 7-kg samples, twenty- 
four 20-L plastic buckets were used. Each bucket 
had two holes drilled in the bottom and was 



110 



Table 1. — Sulfur content of waste materials. 





Sulfur form, pet by weight 


Material 


Total 


Pyritic 


Sulfate 


Organic 


Slightly weathered 












sulfide tailings 


20.6 




19.7 


0.84 


- 


Extensively weathered 












sulfide tailings 


5.98 




0.17 


5.81 


- 


Waste rock 


4.07 




4.07 


- 


- 



Table 2. — Treatment of samples for each waste material. 



Sample weight, 


Chemical 


Treatment 


Volume of water used 


kg 


treatment 


dosage, mg/kg 


in weekly leachings, L 


100 


Untreated 


_ 


8.0 


100 


SLS 


600 


8.0 


100 


Benzoate 


600 


8.0 


7 


Untreated 


- 


1 .2 


7 


it 


- 


1 .2 


7 


SLS 


600 


1 .2 


7 


ii 


60 


1.2 


7 


Benzoate 


600 


1 .2 


7 


n 


60 


1 .2 


7 


Sorbate 


600 


1 .2 


7 


it 


60 


1 .2 



positioned over another bucket to enable leachate 
collection. 

Each of the mine wastes was used to fill three 
barrels and eight buckets with 100 kg and 7 kg of 
material, respectively. The waste materials were 
added (without compaction) to their respective 
containers, and within one week, the experiment 
began. An enclosed hanger-type building in which 
the air temperature ranged from approximately 15 to 
25° C was used for the experiment. 

Each barrel (100-kg samples) designated to 
receive chemical treatment was treated with 8 L of 
a 7,500-mg/L solution of SLS or benzoate. These 
treatments were equivalent to 600 mg of treatment 
chemical per kilogram of waste material, however, 
any treatment solution that did not adsorb to the 
waste material was allowed to drain freely from 
each barrel. The barrels designated as controls 
received 8 L of deionized tap water. 

Two of the eight buckets (7-kg samples) of 
each material were designated as untreated 
controls. The remaining six buckets of each 
material received a high or low dose of SLS, 
potassium benzoate, or potassium sorbate. The high 
treatment dose was added to each bucket using 1.2 L 
of the appropriate 3,500-mg/L solution. The low 
treatment dose was applied using 1.2 L of a 
350-mg/L solution. The high and low treatment 
doses were equivalent to 600 and 60 mg, respec- 
tively, of treatment chemical per kilogram of 
material. The buckets designated as controls 
received 1 .2 L of deionized tap water. 

The weekly leaching program began one week 
after treatment. Tap water was passed through a 
0.20-um capsule filter and a two-bed 
ion-exchange-resin system before storage in a 
500-gal plastic reservoir. Periodic sampling of 
this reservoir showed the water to be acceptably 



low in dissolved solids, and the population of 
iron-oxidizing bacteria was below detectable limits 
(<2 cells/100 mL) . A plastic sprinkling can was 
used to distribute water evenly on each waste 
material. The 100-kg and 7-kg samples were leached 
with 8.0 and 1 . 2 L of water, respectively. These 
volumes are equivalent to 2.5 cm of precipitation 
on the waste material per week. 

After the waste materials were watered, they 
freely drained into the sample-collection buckets 
for 24 hours. Samples from each leachate were 
taken and the remaining volume was measured and 
discarded. Water samples of the leachate from each 
bucket and barrel were analyzed for pH, acidity, 
ferrous iron, total iron, calcium, magnesium, 
aluminum, sodium, manganese, and sulfate. Ferrous 
iron was measured by potassium dichromate titration 
using the sodium salt of di phenyl ami nesulfonic acid 
as an indicator. Other metal concentrations were 
determined by inductively coupled argon plasma 
(ICAP) spectroscopy. After passing the sample 
through a cation-exchange resin, sulfate was 
analyzed by barium chloride titration using thorin 
as an indicator. Acidity analysis consisted of a 
fixed endpoint titration to pH 8.3 using a 1N 
sodium hydroxide solution and converted to mg/L as 
CaC03- 

Analyses for iron-oxidizing bacteria were 
performed for the 100-kg samples (barrels) only. 
The multiple-tube, most probable number (MPN) 
method (American Public Health Association 1985) 
using the media described by Cobley and Haddock 
(1975) was used. This method uses a series of 10- 
fold sample dilutions that are added to a low-pH, 
ferrous-iron-enriched media. Bacterial densities 
are estimated by observing a color change in the 
tubes, which indicates oxidation of ferrous iron by 
iron-oxidizing bacteria. 



Ill 



The percent of iron in the ferrous (as opposed 
to ferric) state in the raw samples was found to be 
a good indicator for the density of iron-oxidizing 
bacteria. Higher ferrous iron percentages 
indicated that the bacterial density was low. 
Conversely, lower ferrous iron percentages 
indicated higher numbers of bacteria. Therefore, 
chemical analysis of the leachate (for ferrous and 
total iron) enabled a rough estimate of the number 
of iron-oxidizing bacteria in the 7-kg samples 
without actual MPN analysis. Figure 1 shows these 
correlations for the 100-kg samples of waste rock 
and slightly weathered sulfide tailings that were 
treated with SLS. 



RESULTS 

The following sections cover the results of 
the three treatment chemicals on the three mine 
waste materials. Although direct measurements of 
the bacterial populations were made, the effec- 
tiveness of the chemicals was based on leachate 
water quality. Graphs of acidity concentration and 
cumulative acid load over time are presented 
because acidity directly relates to water treatment 
costs and acidity was correlated with other water 
quality parameters of interest (table 3). All 
acidity measurements are expressed in CaCG^ 
equivalence. 



Table 3. --Spearman's correlation coefficients* of 
acidity with other water quality parameters. 



Sulfate 
Iron 

Aluminum 
Manganese 
_pH_ 



0.97 
0.94 
0.94 
0.89 
•0.78 




I0 6 
I0 5 
I0 4 



CE 
Id 

I- 

% 

CO. 

2* - 

M-55 

X I0 2 - 

o 

I 

I !0> 



1 1 1 r 

(a) Waste rock 



-o—oii-o- 



20 40 60 80 
PERCENT FERROUS IRON 



100 








1 1 1 






°0 


□ 


(b) Slightly weathered 
sulfide tailings 




- 


a 










- 




o d a 




" 


- 




a 

D 




- 




i 


o 
a a 

1 La o-oi 


D 





40 50 60 70 80 
PERCENT FERROUS IRON 



90 



k All significant at p < 0.001 



Figure 1. --Relationship between the population of 
iron-oxidizing bacteria and percent ferrous 
iron in 100-kg, SLS-treated samples of (a) 
waste rock and (b) slightly weathered sulfide 
tailings. 



Waste Rock 



In the 100-kg samples, the iron-oxidizing 
bacteria were reduced below detectable limits by 
SLS and benzoate for 182 and 231 days, 
respectively (fig. 2) . Both SLS and benzoate 
significantly reduced acidity levels in the 
leachate (fig. 3). After 250 days, the cumulative 
acid load generated by each treated sample was 15 
percent of the cumulative acid load generated by 
the control (fig. 4). Forty days after the 
bacterial inhibitors began to lose effectiveness, 
the iron-oxidizing bacteria repopulated to levels 
greater than those found in the control. The 
concentration of acidity from each treated sample 
also increased above the control level. These 
increases in acidity occurred about 100 days after 
the increases in iron-oxidizing bacteria. 

In the 7-kg samples, the high dose (600 mg/kg) 
of SLS, benzoate and sorbate kept acidity 
concentrations below 50 pet of the lowest of the 
two untreated controls for 343, 117, and 147 days 
respectively (fig. 5a). The low dose (60 mg/kg) of 
SLS and benzoate reduced acidity concentrations 
below 50 pet of the lowest control for 63 and 77 
days, respectively. The low sorbate dose did not 
reduce acidity levels (fig. 5b). 




IOO 



200 
TIME, days 



300 



400 



Figure 2. --Population of iron-oxidizing bacteria 
vs. time for 100-kg samples of waste rock. 



112 



12,000 
10,000 

_j 8,000 

E 

>-* 6,000 

5 

^ 4,000 

2,000 





1 


1 1 1 
KEY 


1 


1 


1 





- Control 






" 


1 — 


- Benzoate 










-SLS 






" 


" \ /* 






/ 




i 

I 






/ 
/ 
/ 


./- 


\ 




S 
• 


/ I 

s • 

' / 


" 


>_ ^ 


•— ISJZJT.Z L 


II 1 ~~ 


-< 


1 



50 100 150 200 250 300 350 
TIME, days 



Figure 3. --Acidity concentration vs. time for 
100-kg samples of waste rock. 



2,000 



§ 1,500 

o 

_i 



^ 1,000 

UJ 

> 



< 

_l 



500 



1 


— r i 

KEY 

- Control 

- Benzoate 
-SLS 


i 


—r 


1 


- 


-T--I-- 


- -t 1 


i 


s 
s 

_l 


/ 
s 

1 


/ 



E 



* 



50 100 150 200 250 300 350 
TIME, days 



E 
Q 



5,000 
4,000 
3,000 r 
2,000 
1,000 

5,000 
4,000 
3,000 f 



1 
(a) 


i 


i 


1 1 1 

KEY 

Control - 

Control 








Benzoate 








SLS 


w 






Sorbate 


1 






i 


\, 


,/-' 


^^^^ 


i 


i_. 


'••!••■ 





50 100 150 200 250 300 350 
TIME, days 

— i r 

(b) 



T 



1 r 




J L 



50 1 00 I 50 200 250 300 350 
TIME, days 

Figure 5. --Acidity concentration vs. time for 

(a) high treatment dose (600 mg/L) and 

(b) low treatment dose (60 mg/kg) on 7-kg 
samples of waste rock. 



Figure 4. --Cumulative acid load vs. time for 
100-kg samples of waste rock. 



Extensively Weathered Sulfide Tailings 

In the 100-kg samples, greater concentrations 
of acidity were leached from the SLS- and 
benzoate-treated samples than from the control in 
the first 40 to 56 days (fig. 6). Bacterial 
populations were not inhibited, and ranged from 
3 x 103 to 2 x 10 5 cells/100 mL. 

Similarly, results from the 7-kg samples 
showed that all of the treatments increased acidity 
over control levels in the first 28 to 56 days of 
the experiment with no subsequent difference in 
acidity levels between treated samples and controls 
for the remainder of the experiment. Ferrous iron 
data suggested that no significant bacteria 
inhibition occurred. 



80,000 



60,000 



KEY. 

Control 

Benzoate 

SLS 




4= 



50 1 00 

TIME, days 



I 50 



Figure 6. --Acidity concentration vs. time for 
100-kg samples of extensively weathered 
sulfide tailings. 



113 



Slightly Weathered Sulfide Tailings 



In the 100-kg samples, neither SLS nor 
potassium benzoate reduced acidity levels relative 
to the control in the leachate from the slightly 
weathered sulfide tailings. However, both 
treatments kept the bacterial population 
significantly below that of the control for 
approximately 1 30 days (fig. 7). In the 7-kg 
samples, none of the treatments had any significant 
effect on leachate water quality. However, based 
on the high percentages of total iron remaining in 
the ferrous form, the iron-oxidizing bacteria were 
inhibited for over 100 days. 

An additional test was performed to determine 
why the inhibition of bacteria in the slightly 
weathered sulfide tailings did not result in lower 
acidity levels. In this test, the tailings were 
washed with a dilute hydrochloric acid solution to 
remove any oxidized products before the treatments 
were added. The experiment used six 7-kg samples, 
three of which were not treated with bacterial 
inhibitor to act as controls. Two samples were 
treated once with a 600 mg/kg dose of SLS and one 
sample was treated once with a 600 mg/kg dose of 
benzoate. Procedures used for leaching and water 
quality analyses were the same as outlined in the 
experimental procedure section. 

Both SLS and benzoate significantly reduced 
acidity below control levels (fig. 8). At the end 
of 150 days, the cumulative acid load generated by 
each treated sample was 18 percent of the acid 
loads produced by the control samples (fig. 9). 
Analysis for iron-oxidizing bacteria showed that 
populations in the SLS- and benzoate-treated 
samples remained below 2 cells/100 mL for 112 and 
98 days, respectively. The population of iron- 
oxidizing bacteria in the control samples ranged 
from 2.3 x 10 4 to 1.3 x 10 5 cells/100 mL. 



1 1 

< 10 

K 9. 

So 7 

o2 6 
z^ 




i 


i 


'O ' 


1 i 
i i 





IRON-OXIDIZI 
log (cells 


KEY 
o Control 
+ Benzoate 
oSLS 

i , i 




40 




80 


120 
TIME, 


1 60 200 240 
days 



Figure 7. --Population of iron-oxidizing bacteria 
vs. time for 100-kg samples of slightly 
weathered sulfide tailings. 




20 40 60 80 100 120 140 160 180 200 
TIME, days 



DISCUSSION 

The leaching procedure used in this experiment 
represents an acceleration of actual field rainfall 
conditions. Although 2.5 cm/week of precipitation 
does not seem excessive, the procedure does not 
allow for runoff or evaporation that might occur in 
the field. The duration of bacterial inhibition 
for the treatment chemicals may have been 
diminished by the short period of time that the 
water was added. This may have increased the 
amount of chemical that was washed off; however, 
the treatment was not lost to runoff and may have 
been deposited at a lower position in the sample. 
Because of these variables, caution must be 
exercised during extrapolation of the results of 
laboratory leaching tests to intended field use. 

Results from the benzoate- and SLS-treatments 
of the waste rock indicate that both chemicals may 
be applicable for field use. SLS has reportedly 
been applied to waste rock by a mining company in 
the field, but no reductions in contaminant levels 
were observed. The blocky nature of the waste rock 
permits air to penetrate deeply into the pile. 
Consequently, in order to inhibit bacteria and 
reduce contaminant levels, a large portion of the 
pile must receive the chemical inhibitor and not 
just the uppermost layer. 



Figure 8. --Acidity concentration vs. time for 
7-kg samples of hydrochloric-acid-rinsed 
sulfide tailings. 



Ui 

> 



300 
250 
200 
150 
100 - 
50 - 




KEY 
» o Control 

* Benzoate 
DA SLS 




20 40 60 80 100 120 140 160 180 200 
TIME, days 

Figure 9. --Cumulative acid load vs. time for 
7-kg samples of hydrochloric-acid-rinsed 
sulfide tailings. 



114 



Results from the extensively weathered sulfide 
tailings showed that all of the treatments 
increased acidity levels in 100- and 7-kg 
samples. The material was considerably oxidized 
with 5.81 pet sulfate sulfur and only 0.17 pet 
pyritic sulfur; therefore, most of the contami- 
nation in the leachate was from the flushing of 
previously oxidized material. Apparently, the 
treatment chemicals caused an increased flushing of 
contaminants. SLS could have increased 
contaminant flushing by its surfactant properties. 
Benzoate and sorbate readily form complexes with 
iron which could result in dissolution of 
weathered products, thereby increasing contaminant 
flushing. 

The results from the -slightly weathered 
sulfide tailings showed that SLS, benzoate, and 
sorbate effectively inhibited iron-oxidizing 
bacteria but did not reduce contaminant levels. 
However, in a subsequent test in which the tailings 
were washed with dilute HC1 to remove any oxidized 
products prior to treatment, SLS and benzoate 
inhibited the bacteria and reduced acidity levels. 
These results indicate that these chemicals may be 
applicable to field use if the tailings can be 
treated before they begin to oxidize. 

In all three materials, differences between 
the results from the 100-kg samples and the 
7-kg samples were noted. In the 100-kg samples of 
extensively weathered sulfide tailings, SLS and 
benzoate initially increased acidity to 28,000 and 
67,000 mg/L, but the leachate from the control 
sample only contained 12,000 mg/L; however, in the 
7-kg samples initial acidity levels were only 
18,000 mg/L for the treatments compared to 10,000 
and 15,000 mg/L for the two controls. Likewise, in 
the 100-kg samples of the slightly weathered 
sulfide tailings, much larger increases in acidity 
levels for the treated samples were observed in the 
larger scale test than in the 7-kg samples. Also, 
the magnitudes of acidity concentration were much 
larger from the 100-kg samples (up to 150,000 mg/L) 
than from the 7-kg samples (below 25,000 mg/L). In 
the waste rock, using the 600 mg/kg dose of 
treatments, the 100-kg samples showed benzoate to 
be slightly more effective than SLS, but in the 7- 
kg samples, SLS was much more effective than 
benzoate. Again, magnitudes of contaminant 
concentration differed between the two sample 
sizes. Although basic trends were fairly 
consistent between the 7- and 100-kg sample 
results, large discrepancies in contaminant 
concentration were observed even under these 
controlled conditions. Therefore, trying to use 
this laboratory data to extrapolate to anticipate 
field results (viz. correct chemical treatment 
dose, duration of effectiveness, reduction of 
acidity and metals levels, and corresponding 
reduction in water treatment costs) cannot be 
accurately made. However, these methods can be 
used to screen potentially effective chemicals and 
provide some guidance for larger scale field 
studies . 



CONCLUSIONS 

The amount of weathering a mine waste has 
undergone prior to treatment with SLS, potassium 
benzoate, or potassium sorbate determines the 
effectiveness of these chemicals to inhibit iron- 
oxidizing bacteria. Application of these chemicals 



on weathered mine wastes can increase the 
liberation of previously oxidized contaminants. 
The presence of even a small amount of weathered 
products on mine wastes can render these 
treatments ineffective. However, if the material 
is unweathered (all sulfur in the pyritic form), 
these treatments can inhibit iron-oxidizing 
bacteria and significantly reduce acidity, sulfate, 
and metal concentrations in the leachate. 

In our tests using sulfide tailings, SLS, 
potassium benzoate, and potassium sorbate did not 
reduce contaminant concentrations in the leachate 
from the any of the weathered tailings. However, 
following removal of the weathered products from 
the slightly weathered sulfide tailings, SLS and 
benzoate effectively reduced contaminant 
concentrations in the leachate. Respectively, SLS 
and benzoate reduced the density of iron-oxidizing 
bacteria below detectable limits (2 cells/mL) for 
100 and 86 days, and reduced the cumulative acid 
load to 18 percent of the control acid load after 
150 days. 

Results of the tests using 100-kg samples of 
waste rock showed that SLS and potassium benzoate 
inhibited the iron-oxidizing bacteria below 2 
cells/100 mL for 182 and 231 days, respectively. 
In these samples, SLS and benzoate reduced the 
cumulative acid load to 15 percent of the acid 
load from the control after 250 days. 



LITERATURE CITED 

American Society for Testing and Materials. 1983. 
Annual book of ASTM standards, v. 05.05 
Gaseous Fuels; Coal and Coke, pp. 347-349. 

APHA. 1985. Standard methods for the examination 
of water and wastewater. 16th Edition, 
pp. 870-886. American Public Health 
Association, Inc., Washington, DC. 

Browning, J. E. 1970. Freshening acid mine- 
waters. Chemical Engineering. 77:40-42. 

Cobley, J. G. and B. A. Haddock. 1975. The 
respiratory chain of Thiobacillus 
f errooxidans : the reduction of cytochromes by 
Fe^ + and the preliminary characterization of 
rusticyanin a novel 'blue' protein. FEBS 
Letters. 60:29-33. 

Cruess, W. V. and P. H. Richert . 1929. Effect of 
hydrogen ion concentration on the toxicity of 
sodium benzoate to microorganisms. Journal of 
Bacteriology. 17:363-371. 

Eklund, T. 1980. Inhibition of growth and uptake 
processes in bacteria by some chemical food 
preservatives. Journal of Applied 
Bacteriology. 48:423-432. 

Eklund, T. 1983. The antimicrobial effect of 

dissociated and undissociated sorbic acid at 
different pH levels. Journal of Applied 
Bacteriology. 54:383-389. 

Kleinmann, R. L. P., D. A. Crerar, and R. R. 

Pacelli. 1980. Bi.ogeochemistry of acid mine 
drainage and a method to control acid 
formation. Mining Engineering. 33:300-306. 



115 



Kleinmann, R. L. P. and P. M. Erickson. 1982. 
Full-scale field trials of a bactericidal 
treatment to control acid mine drainage. In 
proceedings, 1982 Symposium on Surface Mining, 
Hydrology, Sedimentology, and Reclamation, 
University of Kentucky, Lexington, KY, 
December 6-10, 1982, pp. 617-622. 

Kleinmann, R. L. P. and P. M. Erickson. 1983. 
Control of acid drainage from coal refuse 
using anionic surfactants. 16 pp. Bureau of 
Mines Report of Investigation 8847. 

Margaritis, A. and E. Creese. 1979. Toxicity of 
surfactants in the aquatic environment: a 
review. In Waste Treatment and Utilization. 
M. Moo-Young and G.J. Farquhar (Eds.). 
Permagon Press, Oxford, U.K. pp. 145-462. 

Onysko, S. J., R. L. P. Kleinmann, and 

P. M. Erickson. 1984a. Ferrous iron 
oxidation by thiobacillus f errooxi- 
dans : inhibition with benzoic acid, sorbic 
acid, and sodium lauryl sulfate. Applied and 
Environmental Microbiology. 48:229-231. 

Onysko, S. J., P. M. Erickson, R. L. P. Kleinmann, 
and M. Hood. 1984b. Control of acid drainage 
from fresh coal refuse: food preservatives as 
economical alternatives to detergents. In 
proceedings, 1984 Symposium of Surface Mining, 
Hydrology, Sedimentology, and Reclamation; 
University of Kentucky, Lexington, KY, 
December 2-7, 1984, pp. 35-42. 

Onysko, S. J. Chemical abatement of acid mine 
drainage formation. Ph.D. Dissertation, 
University of CA, Berkeley, CA 1985, 315 pp. 

Singer, P. C. and W. Stumm. 1970. Acidic mine 

drainage: the rate-determining step. Science. 

167:1121-1123. 

Sofos, J. N. and F. F. Busta. 1981. Antimicrobial 
activity of sorbate. Journal of Food 
Protection. 44:614-622. 

Stancel , W. J. 1982. Prevention of acid drainage 
from coal refuse. In proceedings, 1982 
Symposium on Surface Mining, Hydrology, 
Sedimentology, and Reclamation, University of 
Kentucky, Lexington, KY , December 6-10, 
1982. pp. 119-122. 

U. S. Code of Federal Regulations. 1983. Title 
21 : Food and Drugs . 



116 



MAPPING BURIED TIPPLE REFUSE - IS THE MAGNETOMETER BETTER THAN 

TERRAIN CONDUCTIVITY? 1 



Joseph H. Schueck, P.E, 



Abstract. — It is occasionally desirable to detect 
and map tipple refuse buried beneath strip mines . 
Reasons include: acid mine drainage (AMD) treatment 
and/or abatement, and economic recovery for 
reprocessing. Until now terrain conductivity (EM) 
has been the most efficient, cost effective method of 
mapping refuse. However, it is difficult to 
delineate between refuse and other features such as 
AMD using EM. Remanent magnetism, a property 
somewhat 'unique' to refuse in the strip mine 
environment allows refuse to be mapped independently 
of other features using a magnetometer. Two sites 
were mapped using both EM and the magnetometer. 
Refuse disposal completely filled . large pits at the 
first site but was limited to widely scattered, small 
pods at the second. When the proper grid density was 
used, the EM was able to define general disposal 
areas but was unable to pinpoint refuse locations, 
especially on the site where the refuse was in 
scattered pods. The EM was unable to discriminate 
between the refuse and AMD. On the other hand, the 
magnetometer was able to clearly define the refuse 
limits on both sites without interference from the 
AMD. Acquisition of data is seven times faster with 
the magnetometer and the magnetometer data provides 
much better quantitative information about the refuse 
than does the EM data. 



INTRODUCTION 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and Reclamation 
and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), April 
17-22, 1988, Pittsburgh, Pa. 

2 
Joseph H. Schueck, P.E., is a 

Hydrogeologist , Pennsylvania Department of 

Environmental Resources, Bureau of Mining 

and Reclamation, Harrisburg, Pa. 



Many advances have been made in 
recent years toward preventing, 
mitigating, abating, and treating acid 
mine drainage. Research efforts have 
concentrated on the source of the acid 
mine drainage as well as the discharge. 
Buried tipple refuse is a source of many 
acid mine drainage discharges. However, 
treatment at or elimination of the source 
cannot be accomplished until the refuse 
has been delineated. Additionally, refuse 
from certain older plants may be rich 
enough to be reprocessed, provided it can 
be found and mapped efficiently. This 
paper discusses a rapid and efficient 
method to accurately map buried tipple 
refuse . 



117 



Tipple refuse is commonly disposed 
of at both active and abandoned strip 
mines. Pits may be completely filled with 
refuse at abandoned sites, and the refuse 
'pods' or buried piles may be quite large. 
At active sites burial locations are 
dictated by the location of the working 
pit. Widely scattered pods of limited 
aerial extent commonly result. Following 
reclamation, there is little evidence to 
indicate the locations of the buried 
refuse. Producing accurate maps of these 
locations has been difficult until now. 

Subsurface features on strip mines 
have been successfully mapped in recent 
years using geophysical techniques such as 
DC electrical resistivity and 
electromagnetic terrain conductivity (EM). 
Refuse usually responds well to both 
methods if it is an electrical conductor. 
However, features such as subsurface 
accumulations of acid mine drainage also 
conduct electricity well and may produce 
anomalies similar to refuse. This 
property makes delineation of refuse 
difficult using either method. 

Another property of at least some 
refuse, magnetism, allows the refuse to be 
detected with a magnetometer. 
Magnetometer operation requires only one 
person. It will be shown that buried 
refuse can be mapped in greater detail and 
several times faster with a magnetometer 
than with EM or DC resistivity. 

Two sites, referred to as SW1 and 
NWi , were mapped during this study. 
Tipple refuse was buried in abandoned 
strip mine pits at SW1 . The resultant pods 
of refuse were up to 700 ft long. Refuse 
was buried in scattered pods controlled by 
the location of active pits on NWI . The 
majority of these pods probably are no 
larger than one or two truckloads of 
refuse. Magnetic detection of the buried 
refuse at both sites was quite successful. 



PRINCIPLES OF MAGNETISM 

The earth's magnetic field resembles 
the field of a large bar magnet near its 
center. Flux lines of the earth's field 
exhibit a pattern like that of a bar 
magnet, figure 1. The field intensity is 
a function of the flux line density. 
Similar to a bar magnet, the density is 
twice as large in the polar regions as in 
the equatorial regions, or approximately 
60,000 and 30,000 gammas respectively. 
The total field is approximately 55,500 
gammas in Pennsylvania. By convention in 
the geophysical community, gauss is the 
cgs unit,, of magnetic intensity and 1 
gamma=10 " gauss. The earth's total 
magnetic field intensity is measured using 
a magnetometer (Breiner 1973). Most 
commercially-available magnetometers 
measure the magnitude of the total field 
intensity. Older, "f luxgate"-type, 
magnetometers are sometimes seen. These 
measure components of the total field 
intensity vector. 







Figure 1. The earth's magnetic field 
resembles the field of a large bar magnet . 
The flux line density is greater at the 
polar regions than at the equator (Breiner 
1973). 



An anomaly represents a local 
disturbance in the earth's magnetic field 
due to a local change in magnetization. 
Magnetic minerals and iron have properties 
which cause these anomalies . Natural 
anomalies are due chiefly to the presence 
of magnetite (FeFe-O.), or its related 
mineral suite: ulvospinel, maghemite, etc. 
All rocks contain some magnetite from a 
very small fraction of a percent up to 
several percent. Iron objects produce 
strong, local magnetic anomalies (Breiner 
1973) . 

Magnetic anomalies are caused by two 
different kinds of magnetism: induced and 
remanent or "permanent" magnetization. 
Induced magnetization refers to the effect 
on a material placed in an external 
magnetic field in which the field within 
the material is enhanced, so that the 
material itself acts as a magnet. Induced 
magnetization is always parallel to the 
inducing field (usually the earth's), and 
would vanish if this inducing field were 
removed. The magnetization of such 
material is directly proportional to the 
intensity of the ambient field and to the 
ability of the material to enhance the 
local field. This property is called 
magnetic susceptibility. The induced 
magnetization is equal to I.=kF, where I. 
is the induced magnetization per unil; 
volume in cgs electromagnetic units, k is 
the volume magnetic susceptibility, and F 
is the field intensity in gauss. k is 
usually between 10 and 10 cgs or for 
most rocks. For pure magnetite, k is 
approximately . 3 cgs . k may vary between 
1 and 10 for iron alloys (Breiner 1973) 



118 



Materials cooling from high 
temperatures may attain a remanent 
magnetization. This remanent 
magnetization, once gained, is independent 
of external fields.. it need not be 
parallel to the external field and will 
remain even if the external field is 
removed. Remanent magnetization, I r , is 
often ten or more times greater than 
induced magnetization in many igneous 
rocks and iron alloys . Permanent 
magnetization depends upon the 
metallurgical properties and the 
mechanical, thermal, and magnetic history 
of the specimen. Prior to heating, small 
regions, called domains, within each 
magnetite crystal are randomly oriented. 
The domains reorient themselves during 
heating. Upon cooling, they align 
parallel to each other in the direction of 
the ambient magnetic field. This creates 
a net magnetization fixed with respect to 
orientation of the object. Magnetite may 
have a remanent magnetization, . I , of 
perhaps 0.1 to 1.0 gauss (10 to 10 
gammas ) . Ordinary iron may have a 
permanent (.magnetization between 1 and 10 
gauss (10 and 10 gammas). Thus the net 
magnetization might be considerably higher 
and oriented in a different direction than 
would be indicated merely by consideration 
of the susceptibilities (Breiner 1973). 



significant factor in the magnetometer's 
ability to detect refuse. Microscopic 
examination of several samples indicates 
that magnetite is common at SW1 where 
refuse was buried in large pods . 
Combustion products are also present but 
are uncommon. By contrast, at NW1 , 
combustion products are common and 
magnetite crystals are rare. The recorded 
intensities per unit mass of refuse were 
much higher at NW1 where the combustion 
products are predominant than they were at 
SW1 where magnetite is predominant. 



MAGNETIC ANOMALY CHARACTERISTICS 

Magnetic anomalies are highly 
variable in shape and amplitude; they are 
almost always asymmetrical, sometimes 
appear complex even from simple sources, 
and usually portray the combined magnetic 
effects of several sources. An anomaly's 
shape or signature depends upon the 
inclination of the field, and several 
properties of the source: burial depth, 
size and shape of the source, the relative 
amounts of permanent and induced 
magnetization, the direction of the 
former, and the amount of magnetic 
material present in the source compared to 
the adjacent rocks (Breiner 1973). 



SOURCES OF MAGNETISM IN TIPPLE REFUSE 

Samples of refuse and spoil material 
from both sites were analyzed to determine 
the magnetic components unique to the 
refuse. The samples were obtained by air 
rotary drilling. A magnetic separator, 
known as a ferrofilter, was used to remove 
the magnetic portion of the sample. In 
all cases the magnetic portion of the 
sample amounted to only a small fraction 
of a percent. Two magnetic materials were 
identified: magnetite crystals (FeFe 2 0.), 
and reddish grains and black glassy B.B.'s 
that appear to be combustion products 
containing magnetite or maghemite (Fe~0-.) 
unintentionally synthesized during 
combustion (Smith 1987). These particles 
commonly exhibited a 'slag '-like 
appearance and were often fused with 
quartz crystals. The portions of the 
samples which had undergone combustion 
were so minute that they were not detected 
in an unfiltered sample using a 
hand lens. Maghemite forms by slow 
oxidation at low temperatures from 
magnetite or from the slow thermal 
dehydration of lepidocrocite (FeO(OH)) 
below temperatures of 7 50 . The magnetic 
properties of maghemite are similar to 
those of magnetite (Palache, et. al 
1944). Company officials from 
both source tipples indicated that the 
source plants used magnetite in the coal 
separation process at the time the refuse 
was buried on the subject sites. 

An increase in the net magnetism 
caused by the combustion or heating and 
dehydration of the refuse appears to be a 



The dip or inclination of the 
earth's field establishes the direction in 
which the components are measured for any 
local magnetic anomalies using a total 
field magnetometer. The magnetometer only 
measures the component of a local 
disturbance in the direction of the 
earth's field at the point of measurement, 
figure 2. Other factors being equal, the 
same source will produce three different 
anomaly signatures if measured at the 
equator, in Pennsylvania, and near the 
north pole. 

The variation of intensity as 
measured by the magnetometer diminishes as 
the source-to-sensor distance increases. 
Depending on the shape of the source body, 
this fall-off factor usually goes as 1/r 
to 1/r , where r is the distance between 
the source and the sensor. For this 
reason, anomalies will be larger in aerial 
extent than the actual source of the 
anomaly, figure 3. 

The depth to and configuration of 
the source determines the anomaly 
wavelength; the deeper and larger the 
source, the broader the anomaly. Sources 
close to the ground surface exhibit a 
steep gradient between the negative and 
positive portions of the anomaly. Deeper 
sources exhibit a much flatter gradient. 
Figure 4 shows two anomalies which were 
mapped . The depth to the source on the 
left is about 12.5 ft compared to a 50 ft 
burial depth on the right. 

The negative part of an anomaly is 
called its "polarity low" . At the 
latitude of Pennsylvania, a body whose 



119 




-w V \1 Xx 



^'ST MONOPOLE 



Figure 2. The portable proton magnetometer only measures the component of a 
local disturbance in the direction of the earth's field at the point of 
measurement (Breiner 1973). 





DIPOLE 



MONOPOLE 




Figure 3. The variation of intensity as measured by the magnetometer diminishes as 
the source-to-sensor distance increases. The fall-off factor usually goes as 1/r 3 
for dipoles to 1/r 2 for monopoles. Resultant anomaly signatures are shown at the 
bottom of the figure for dipoles, left, and monopoles, right (Breiner 1973). 



magnetization is mostly induced will 
usually have a small polarity low to the 
magnetic north, with the maximum part of 
its high over the southern part of the 
body. A body with very strong remanent 
magnetization, by contrast, will have an 
adjacent polarity low in some direction 
other than north. If such a pattern is 
seen on a magnetic map, it usually 
indicates sources with strong remanence. 

Maximum anomaly amplitude depends 
upon several factors. These include 
depth, the contrast in the mass of 
magnetic material to the surrounding 
material, and the shape and configuration 
of the source. Small, compact sources will 
appear as a dipole, i.e. the anomaly 
signature will show both an anomaly high 
and a polarity low. If the source is 
narrow, broad, or long in one dimension, 
the anomaly may appear as a monopole, i.e. 



only a positive portion may be recorded or 
the polarity low may appear beyond one of 
the ends . 



FIELD PROCEDURES 

A portable proton magnetometer was 
used at both sites . This instrument 
measures total field intensity only. It 
has a sensitivity of + or - 1 gamma over a 
range from 20,000 to 90,000 gammas and a 
gradient tolerance which exceeds 800 
gammas/ft. The instrument consists of a 
sensor mounted on a 8-ft long staff 
connected to a power pack and digital 
readout display (Geometries). 

A grid network along with a base 
station or baseline should be established 
over the site before readings are taken. 
Grid spacing depends on the anticipated 



120 




o 



/ 



O 
0} 



13^t "--. --' .. 



:aie 



Anomaly High 

Polarity Low 

Contour Interval = 20 Gammas 




Anomaly High 



Polarity Low 

Contour Interval = 5 Gammas 



Figure 4. The depth to and configuration of the source body controls the anomaly 
signature. The deeper and larger the source body, the broader the anomaly. 
Shallow sources exhibit a steep gradient between the anomaly high and the polarity 
low; deep sources produce a flat gradient. The depth to the source body on the 
left is 12.5 ft compared to a 50 ft burial depth on the right. Note scale and 
contour interval differences. 



target size. Readings are taken along the 
grid intersections with frequent 
measurements made at the base station or 
along the base line. 

Magnetometer operation is quite 
simple. The staff with the sensor mounted 
on top is placed on the ground at the 
point where the reading is to be taken and 
is held still. A read button is pushed. 
Seconds later a digital readout appears 
which is recorded. The sequence is 
repeated at each grid intersection. Data 
acquisition is quite rapid. However, a 
few precautions must be taken in order to 
insure meaningful data. 

Quality control is a key to 
obtaining reliable field data. A proper 
grid density must be selected, natural 
variations in the earth's field intensity 
must be tracked, and local magnetic 
anomalies not associated with the target 
must be located and noted. 

The aerial extent of the refuse 
dictates the maximum grid spacing or 
station density which may be used. Large 
pods on the order of tens or hundreds of 
feet can be adequately mapped on a 25 by 
50 ft grid. Truck load sized pods of 
refuse can be located with a 25 by 25 ft 
grid; however, reasonable definition of 
the refuse requires returning to the 
target area and taking additional readings 
on a tighter grid. 

Recording variations in the earth's 
total field with time and removing these 
variations from the field data are crucial 
to quality data. Magnetic storms and 
micropulsations due to solar winds acting 



on the ionosphere cause variations of 10 's 
to 100 's of gammas. These variations are 
not predictable. They may occur over a 
short period of 10 's of minutes and be 
quite irregular or they may occur slowly 
over a period of a day (Breiner 1973). A 
recording base station magnetometer on or 
near the site can record these variations 
while field readings are being taken 
elsewhere. The base station readings are 
then used to correct the field readings. 
Otherwise it is necessary to establish a 
base station or base line and return to 
the station frequently, usually after 
every traverse, to record the variations. 
Figure 5 shows a variation of nearly 35 
gammas which occurred at the base station 
on SW1 over an 8-hour time span. Daily 
variations such as this were common at 
both sites. 




2 4 6 
TIME IN HOURS 

Figure 5. Jormal Daily Magnetic Fluctuations 



121 



Scrap metal and other iron objects 
lying on the surface must be removed to an 
area beyond their influence on the 
readings. Locations of metal objects 
which cannot be removed from the area, 
such as well casings, should be noted and 
their influence taken into consideration 
during the interpretation. Examples of 
such interferences follow. At NW1 small 
plastic flags on wires were used to mark 
grid locations . Readings obtained 
adjacent to these flags resulted in 10 to 
15 gamma anomalies. A metal well casing 
and a galvanized garbage can produced 225 
and 17 gamma anomalies, respectively, 
with the magnetometer adjacent to the 
metal. The anomalous reading diminished 
to at about 20 ft in both cases. 
A drag line cable buried just beneath the 
surface caused a 10,000 gamma anomaly on 
another site. 



FIELD STUDY AT SW1 SITE 

Site Description 

SW1 is a 43 acre reclaimed refuse 
disposal site in Somerset County, Pa. 
Once an active strip mine, the original 
operator abandoned the site after coal 
removal, leaving behind several 
unreclaimed pits and spoil piles. The 
pits were as deep as 60 ft and several 
were longer than 500 ft. Pit width varied 
from 50 to 200 ft. Some pits had vertical 
unreclaimed highwalls while others had 
sideslopes of spoil. 

All major pits on the site were 
later permitted for tipple refuse 
disposal. Inspection reports and 
subsequent drill holes indicate that the 
operator completely filled the pits with 
The site was then regraded and 
by this operator, leaving no 
evidence of the buried refuse 



refuse, 
reclaimed 
surf icial 
locations , 



Conductivity Field Work 

The southern half of SW1 was mapped 
in 1986 using electromagnetic terrain 
conductivity. The purpose was to 
determine the refuse limits so that 
monitoring wells could be strategically 
located. The area was mapped on a 50 by 
50 ft grid and a total of 191 readings 
were made. A 10-m intercoil spacing with 
a horizontal coplanar dipole configuration 
was used. It was known that burial of the 
refuse extended to near the surface. This 
configuration has an effective exploration 
depth of 7.5 m and is more responsive to 
the near surface material (McNeill 1980). 



Conductivity Survey Results and 
Interpretation 

A contour map of the EM data is shown 
in figure 6 . Delineation of the refuse 
from the EM data was difficult for two 
reasons. First, the conductivity values 
ranged from a low of 5.5 to a high of 17.5 
mmhos/m. Conductivity measurements over 
refuse are commonly in the 20 to 40 
mmhos/m range on many sites, especially 
considering the thickness and extent of 
the refuse here(Ladwig 1982). Second, a 
rapid increase in the conductivity contour 
gradient was not observed, making boundary 
determinations difficult. This suggests a 
gradual thickening of the refuse, as is 
the case. The contours did steepen 
slightly beyond 8.5 mmhos/m, and it was 
decided that values greater than this 
probably indicated refuse. 

Highs on the EM contour map 
correspond fairly well with outlines of 
the pits scaled from aerial photos taken 
prior to refuse disposal. These 
boundaries compare well with the pit 
limits indicated on the permit maps . The 
maximum conductivity values generally 
coincide with or lie just to the north of 




TEN METER HORIZONTAL DIPOLES 
I CONTOUR INTERVAL .5 mmhos/m 



200 ft NORTH 



Figure 6. Electromagnetic Terrain Conductivity contour map resulting from the 
survey conducted on the southern half of the SW1 coal refuse disposal site. A 
10-m horizontal dipole coplanar configuration was used. Refuse was disposed of 
within the pit boundaries which are indicated. Note the absence of rapid 
contour steepening and conductivity highs beyond the pit boundaries. 



122 



the deepest portions of the pits . Four 
monitoring wells were subsequently drilled 
within areas interpreted to be refuse. 
Refuse depths and thicknesses as indicated 
in the four wells correlate well with the 
■conductivity data. 

There are high conductivity values 
in two areas where very little refuse, if 
any, is thought to exist. The first is to 
the north of the pit in the southwestern 
corner. This area is on the nose of a 
steep pre-disposal spoil pile. The amount 
of refuse in this location should be 
minimal. The second area is to the east 
of the pit in the northeastern portion of 
the study area . Aerial photos indicate 
pre-disposal spoil piles in this area 
also. It is unlikely the operator 
excavated into the spoil to dispose of 
refuse. These values may indicate areas 
of AMD, however. 

A maximum conductivity value of 12 
mmhos/m is observed in the vicinity of 
well P4 . This long narrow pit is filled 
with refuse to a total depth of 60 ft. By 
comparison, the maximum value observed 
over the pit to the east containing wells 
P6 and P5 is 17.5 mmhos/m. The thickness 
of the refuse in this pit is also 60 ft. 
The low value of 12 mmhos/m is misleading, 
but this may be a function of the pit 
geometry. 



Magnetometer Field Work 

SW1 was remapped in June 1987 using 
a portable proton magnetometer borrowed 
from the Pa . Geologic and Topographic 
Survey. The limits of the study were 
extended northward to cover the entire 43 
acres . Readings were initially taken on a 
50 by 25 ft grid. Some areas were 
remapped on a 25 by 25 ft grid for better 
resolution. 

Magnetometer Survey Results and 
Interpretation 

The magnetic anomaly map for SW1 is 
presented in figure 7 . The positive 
anomalies are shown as solid lines. 
Dashed lines indicate the polarity lows . 
The pit boundaries are also indicated as 
in figure 6 . Two basic anomaly signatures 
are repeated: 1) relatively flat gradient, 
maximum 40 gamma anomalies over the long 
pits to the northwest, and 2) relatively 
strong positive anomalies with steep 
gradients and maximum values of about 200 
gammas over the pits in the southwestern 
and eastern portions of the figure. Keep 
in mind that the refuse pods in this 
figure are somewhat smaller than the area 
of the anomalies they cause. Considering 
this, the positive magnetic anomalies 
correlate quite well with the pit 
locations despite the different anomaly 
signatures that are present. 



The anomalies over the northwestern 
pits are consistent with the signatures 
expected over long, narrow dike-like 
structures . The anomaly signature is like 
that of a line of monopoles, i.e. there is 
no distinct polarity low associated with 
the anomaly. Because the magnetometer 
measures the net magnetization of the mass 
within the radius of influence, the 
maximum amplitudes over long, narrow 
structures will be less than that observed 
over wider structures of similar total 
depth . 

The anomaly signatures over the 
remaining pits are also consistent with 
those expected for wide, thick source 
bodies buried at a shallow depth. The 
prominent polarity lows to the north of 
these highs suggest their net 
magnetization is aligned with the earth's 
field, and so may be dominantly of the 
induced kind . The magnetic components of 
the source bodies here are therefore 
inferred to consist largely of magnetite 
rather than combustion products . It is 
further inferred that the magnetite's 
domains of permanent magnetization were 
randomized when the refuse was dumped in 
the pits . The high amplitude of the 
anomaly implies a relatively higher 
concentration of magnetic materials in 
these pits than in those to the northwest. 



The southeastern pit in figure 7 is 
presented in figure 8 to illustrate how 
the observed values correspond to 
thickness of the source body. Lines 
indicating the pit boundary, pit bottom, 
and spoil pile intersections are included. 
Coal pavement was exposed in the western 
portion of the pit prior to filling with 
refuse. A vertical highwall of bedrock 
formed the western pit boundary. The 
eastern portion of the pit was formed by 
spoil piles at their angle of repose 
joining near the center of the pit. The 
haul road into the pit was located in the 
extreme northwestern corner of the pit 
where no boundary line is shown. Refuse 
completely fills this pit to within a few 
feet of the surface. The refuse over the 
pavement in the western portion of the pit 
extends from near surface to a depth of 60 
ft. The depth of refuse in the east-west 
trending portion of the pit with sloping 
sides formed from spoil approaches a 
maximum of 60 ft while the depth of refuse 
in the southeastern portion is less than 
40 ft. The magnetic map, therefore, 
appears to reflect many of the geometric 
details of the original pit. 



Comparison of EM and Magnetometer Results 

Clearly the magnetometer has 
defined the refuse more accurately than 
the EM, both qualitatively and 
quantitatively. The two areas beyond the 
pits in figure 6 which could easily be 



123 












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Figure 8. A enlargement of the pit located in the southeastern portion of SW1, 
figure 7, showing the correspondence between the magnetic contours and the pit 
geometry. The pit boundary is indicated by a solid bold line; the bottom of 
the pit is indicated by a bold dashed line. Prior to refuse disposal the coal 
pavement was exposed in the western portion of the pit and the western highwall 
was vertical. The central and eastern portions of the pit are formed by spoil 
piles joining in the center of the pit at the bottom. 



interpreted as refuse using EM disappear 
when the magnetometer is used. EM could 
define the refuse more clearly had 
additional traverses been completed using 
different intercoil spacings and vertical 
as well as horizontal coplanar 
configurations. However, I was able to 
collect seven times as many data points 
with the magnetometer per man-hour of 
effort than with EM. Furthermore, one 
traverse is all that is needed with the 
magnetometer for quantitative 
interpretations, as opposed to two or four 
with EM. Thus, for good quantitative EM 
data it would be necessary to spend 28 man 
days in the field compared with one with 
the magnetometer. DC electrical 
resistivity profiling could have been used 
to provide electrical information similar 
to EM; note, however, that this technique 
is some three to six times more time 
consuming than EM (Ladwig 1982). 

On the basis of theory, EM 
techniques must be expected to do poorly 
over magnetic targets. The apparent 
conductivity as measured by the 
electromagnetic terrain conductivity meter 
is defined as 



cr. 



H 



.c field at the 



field at the 



w>«/s H 
where H = secondary magnetic 

receiver coil 

H = primary magnetic 

receiver coil 

w = 27ff 
^ = permeability of free space 

cr = ground conductivity (mho/m) 

s = intercoil spacing (m) 

f = frequency (Hz) (McNeill 1980) 
The >( in this equation is usually taken as 
u the permeability of free space under 

tne assumption of nonmagnetic targets. 
For a magnetic target, /-( is higher than 
M , so the conductivity the instrument 
measures comes out lower than the true 
apparent conductivity. The fact that the 
target is magnetic actually degrades the 
EM measurement. 

The mapping effort on SW1 clearly 
demonstrates the effectiveness of mapping 
refuse buried in pods with singular 
dimensions in the tens to hundreds of 
feet. A second site NW1 was mapped to 
determine the effectiveness of the 
magnetometer in mapping refuse buried in 
much smaller, widely scattered pods. 



125 



FIELD STUDY AT NW1 SITE 



Conductivity Field Work 



Site Description 

In contrast to SW1 , refuse disposal 
operations at NW1 occurred while the site 
was actively being mined. NW1 is a 
40-acre site in Clarion County, Pa. The 
extreme western portion of the site was 
mined by a previous operator with perhaps 
only one or two crop line cuts being 
taken. The remainder of 40acre site was 
mined in 1971 and 1972. Reclamation was 
completed in 1973 (Ladwig 1982). No 
particular portion of the site was 
dedicated to refuse disposal; rather the 
refuse was simply dumped into any working 
pit that happened to be open at that time. 
Consequently only one or two truckloads 
make up the entire refuse pod in most 
locations . A typical coal dump truck has a 
capacity of about 25 cu yd. The resultant 
pods or piles would be no more than about 
15-20 ft in the longest dimension. 



A portion of NW1 was mapped in 1982 
using EM by Ken Ladwig, formerly with the 
U.S. Bureau of Mines. Much of the 
remaining portion was mapped in 1984 using 
the same technique. The purpose of the 
mapping was to locate AMD and AMD sources . 
Buried tipple refuse is suspected as one 
of those sources . The EM mapping was 
completed using both 10- and 20-m 
intercoil separations with both horizontal 
and vertical coplanar dipole 
configurations. Readings were taken every 
10 m along traverses approximately 25 m 
apart (Ladwig 1982). 



Conductivity Survey Results and 
Interpretation 

Figure 9 is typical of the 
conductivity contour maps produced from 
NW1 . This figure covers the southeastern 
portion of NW1 . Several large, steep 




Figure 9. A typical terrain conductivity contour map resulting from Ladwig|s 
work at NW1. The area represented is in the southeastern portion of the site. 
Conductivity highs represent AMD or AMD sources (Ladwig, 1982) 



126 



gradient anomalies were mapped at the site 
which Ladwig interpreted as signifying AMD 
or AMD sources. Without drilling, one 
cannot distinguish between anomalies that 
reflect AMD and those that reflect refuse. 
Several confirmation holes were drilled at 
the locations shown in figure 9 . Only 2 
holes, X and F, encountered refuse. 
Ladwig concluded that the anomalies were 
produced by a combination of both AMD and 
refuse (Ladwig 1982). In short, the EM 
work on this site demonstrates the 
similarity in electrical conductance 
properties between AMD and refuse as well 
as the difficulty in identifying either 
one with confidence. 



Magnetometer Field Work 

The entire 40-acre site was mapped 
in July 1987 using a magnetometer. 
Readings were initially taken on a 25 by 
50 ft grid spacing. A contour plot of the 
readings is presented in figure 10. Due 
to resolution, only a 10-acre portion of 
the 40 acres is shown. At this rather 
coarse spacing, many isolated anomalous 
readings were observed. Consequently, to 
get more detail, additional traverse lines 
were added resulting in over half the site 
being mapped on a 25 by 25 ft or tighter 
grid. It should be noted that the area 
presented in figure 10 is located to the 
west of that presented in figure 9. 
Further, EM contours from Ladwig 's second 
effort have been superimposed over figure 
10 for comparison and later discussion. 



Magnetometer Survey Results and 
interpretation 

Many anomaly highs and polarity lows 
are observed in the magnetic anomaly map, 
figure 10. The positive anomalies are 
indicated by solid lines; the polarity 
lows as dashed lines . Both high and low 
anomalies are shaped like circles or 
rounded rectangles. Additionally, maximum 
anomaly amplitudes vary considerably, in 
part because of the grid spacing used. 
Most of the piles can be located, but not 
delineated because only a few data points 
define each anomaly. If the grid point 
where the reading is taken does not 
coincide with the point of maximum 
amplitude for the pile, readings of a few 
tens of gammas rather than a few hundred 
are commonly recorded. Closer grid 
spacing would be required to better define 
these anomalies, but wasn't deemed 
necessary. The 25 by 25 ft grid spacing 
was sufficient to locate the small 
isolated refuse piles. 

Increased resolution through closer 
grid spacing is illustrated in figures 11 
and 12. The anomaly to be redefined is 
the one located in the lower left portion 
of figure 10. The area in figure 11 is 
the result of mapping on a 25 by 25 ft 
grid. The appearance is that of a single 
source with the positive portion of the 



anomaly (solid lines) on the left and the 
polarity low (dashed lines) on the right 
with about a 50 ft separation. The area 
was remapped on a 5 by 5 ft grid as shown 
in figure 12. We find there are two 
sources, each consisting of probably a 
single truckload of refuse. The anomaly 
wavelength indicates a lateral dimension 
of about 15 ft. The steep contour 
gradient further indicates that the burial 
depth is shallow. A drill hole located 
over the anomaly on the right encountered 
refuse from 12.5 to 27.5 ft. 

Additional confirmation of the 
presence of refuse was made through DC 
electrical resistivity soundings using the 
Schlumberger array. Soundings were taken 
over both positive magnetic anomalies as 
well as at locations beyond the anomalies. 
The soundings over the anomalies indicate 
low resistivity (high conductivity) zones. 
Soundings taken away from the anomalies 
indicated uniform, poorly conductive 
materials both near the surface and at 
depth. The presence of refuse beneath 
anomalies was confirmed at several other 
locations across the site using both air 
rotary drilling and DC electrical 
resistivity soundings. In contrast to the 
EM electrical methods, DC electrical 
techniques like Schlumberger soundings are 
not affected by unusual magnetic 
permeabilities. However, standard 
interpretation packages for Schlumberger 
soundings assume infinite flat layers 
rather than the compact local sources we 
know ■ to be present here. Hence, at this 
location the Schlumberger DC results are 
useful quantitatively, but probably not 
qualitatively. 

The presence of polarity lows of 
arbitrary amplitudes that are located in 
an arbitrary direction from the highs, 
figures 10 and 12, indicate that remanent 
magnetization resulting from spontaneous 
combustion is dominant in these source 
bodies . This further indicates that 
spontaneous combustion occurred in the 
refuse before it was trucked onto the 
site. Several facts need to be recalled: 
1) the magnetic fraction of all samples 
from both NW1 and SW1 was minute, but 
comparable in volume 2) magnetic 
combustion particles were predominant at 
NW1, 3) remanent magnetization, a result 
of combustion, is often many times 
stronger than induced magnetization, and 
4) if reoriented, remanent magnetization 
is independent of the local magnetic 
field. First, the size of refuse piles 
are orders of magnitude smaller at NW1 
than at SW1 , yet maximum anomaly 
amplitudes are twice as great at NW1 . 
This indicates greater magnetic 
susceptibilities at NW1 . Lastly, if 
induced magnetization were predominant, or 
if the spontaneous combustion occurred 
after the refuse had been dumped on the 
site, all resulting anomalies would be 
oriented north-south with the polarity 
lows to the north. Such is not the case 
here. Polarity lows do not exhibit a 



127 




Apparent Terrain Conductivity Contour 
20-m Vertical Coplanar Coil Configuration 
2 mmho/m Conductivity Contour Interval 
Magnetic Anomaly High 

Magnetic Polarity Low 

10 Gamma Magnetic Contour Interval 
This area shown in figs 11 and 12 

100 ft 
SCALE i | I NORTH 



Figure 10. Superimposed magnetic and terrain conductivity maps of the SW 10-acres 
of NW1 site. Anomaly highs indicate the location of truckload sized pods of 
refuse buried at the site. Except for a few correspondences, detection of refuse 
by EM methods were poor. (Conductivity contours from Ladwig, 1984). 



uniform north-south alignment. We infer 
from this that the source bodies came off 
the dump trucks in fairly large coherent 
blocks permitting the domains within the 
magnetic components to remain aligned. 



Comparison of EM and Magnetometer Results 

In figure 10 Ladwig 's conductivity 
contours are superimposed over the 
magnetometer contours. Only a 10-acre 



128 




ol lio ft 

Scale 
C.I. = 20 Gammas 



oV-n 







^oiv_-_ 



-«o-. 




Figure 11. Magnetometer contours 
resulting from truckload sized pods 
of refuse mapped on a 25 by 25 ft 
grid. This appears as a single 
source body with the anomaly high to 
the north and the polarity low to the 
south. 



/ tip / / ^'Z-*o\ N \vSS 

i ' ' / ' ' ' i ! 0»" 







I 'A^ 



/rS» 






J j to 



/ 



y 



Figure 12. Magnetometer contour map 
resulting from mapping the area shown 
in Figure 11 on a 5 by 5 ft grid. Note 
the increased detail indicating the 
presence of two anomalies rather than 
just one. 



portion of the 40-acre site is shown. 
Ladwig found a background conductivity 
value of 10 mmhos/m or less at this site 
(Ladwig 1982). An area of values greater 
than 10 mmhos/m is present in the 
left-central portion of the figure, 
implying possible AMD or AMD sources 
there. Note that several magnetic 
anomalies interpreted as refuse locations 
are present in this area. The 16 mmhos/m 
contour coincides with a magnetic anomaly, 
but there appears to be few further 
correspondences between the EM contours 
and the refuse pile locations as 
interpreted magnetically. Also, a number 
of refuse anomalies are mapped using 



magnetics further to the right where EM 
values are below 10 mmhos/m. Also the 
conductivity contours cut across the 
magnetic anomalies rather than surround 
them as would be expected . 

Additional EM readings at much 
closer intervals would probably have 
identified more of the refuse locations. 
However, it may well be that the EM is 
more responsive to the AMD than to the 
refuse at this site, hence the 
conductivity may be showing an AMD pattern 
or the locations of nonmagnetic AMD 
sources . 



129 



CONCLUSIONS 



ACKNOWLEDGMENTS 



The magnetometer's ability to 
clearly define buried refuse regardless of 
pod size makes it an effective geophysical 
tool in locating buried tipple refuse, a 
potential source of acid mine drainage. 
The property of remanent magnetism, which 
is somewhat unique to this particular type 
of refuse, allows the magnetometer to 
identify it against a background of mine 
spoil as well as AMD. 

When attempting to map the locations 
of buried refuse, the magnetometer has 
many advantages over terrain conductivity. 
These include: 

1) The location of the refuse is not 
masked by AMD or highly conductive 
spoil material as it is with EM. 

2) Small piles of refuse, no larger 
than a single truckload are easily 
detected with the magnetometer. 

3) Quantitative as well as 
qualitative determinations can be 
made with magnetometer data obtained 
from a single traverse of the area. 
A single traverse with EM provides 
qualitative data but multiple 
traverses are required before 
quantitative interpretations are 
made. 

4 ) An area can be mapped from 7 to 
28 times faster with the 
magnetometer than with EM for an 
equivalent amount of data. 

5) Quantitative interpretations are 
easier and much quicker with the 
magnetometer than with EM. 



Terrain conductivity has an 
advantage over the magnetometer, however. 
EM can detect AMD as well as structural 
features which the magnetometer cannot. 

It was noted earlier that the refuse 
disposed of at both sites originated from 
tipples where magnetite was used. The 
ability of the magnetometer to detect 
refuse from plants which did not use 
magnetite is as yet unknown. Indications 
are, however, that the refuse may still be 
detected provided some degree of 
spontaneous combustion has taken place. 

Selection of method in a particular 
investigation should be dictated by the 
purpose of the study. If refuse is to be 
mapped for the purpose of reprocessing 
then the magnetometer should be used. 
However, if the study is for delineating 
AMD and its sources, then a combination of 
both methods is suggested. Even though an 
area can be mapped more rapidly with the 
magnetometer, the total field time with 
either method is insignificant considering 
the wealth of information which can be 
collected. 



The author would like to thank 
Robert Smith, II, Chief of the Mineral 
Resources Division, Pa. Geologic and 
Topographic Survey, for his contribution 
in analyzing refuse samples from the two 
study sites. The author would also like 
to thank Dave Campbell, Geophysist with 
the USGS Geophysics Branch in Denver, 
Colorado for his thorough technical review 
and comments . 



REFERENCES 

Breiner, S., 1973. Applications Manual for 
Portable Magnetometer . Geometries , 
Sunnyvale, CA, 57 pp. 

Ladwig, K.J., 1982. Delineation of Zones 
of Acid Mine Drainage Using Surface 
Geophysics in: Proceedings of 1982 
Symposium of Surface Mining, 
Hydrology, Sedimentology, and 
Reclamation, Lexington, KY, pp. 
279-287. 



Ladwig, K.J. 
Manuscript. 



1984, 



Unpublished 



McNeill, J.D., 1980. Electromagnetic 
Terrain Conductivity Measurements at 
Low Induction Numbers . Technical 
Note 6, Geonics Limited, 
Mississauga, Ontario, Canada. 

Palache, C, H. Berman, and C. Frondel, 
1944. The System of Mineralogy of 
James Dwight Dana and Edward 
Salisbury Dana, Volume I. John Wiley 
and Sons, Inc., London, pp. 
643,698,708. 

Smith, R. II, 1987. Magnetic Coal Prep 
Tailings, Interoffice Memo from 
Robert Smith, Pa. Geologic and 
Topographic Survey to Joseph 
Schueck, Bureau of Mining and 
Reclamation, October 26, 1987. 

, Operating Manual, Model 

G-816/826 Portable Proton 
Magnetometer. EG&G Geometries, 
Synnyvale, CA. 12pp. 



130 



THE USE OF PRE-AERATION TO REDUCE THE COST 
OF NEUTRALIZING ACID MINE DRAINAGE 1 



T. C. Jageman, R. A. Yokley and G. W. Heunisch 2 

Consolidation Coal Company 

Research and Development Department 

4000 Brownsville Road 

Library, Pennsylvania 15129 



Abstract. — Acid mine drainage (AMD) pumped from 
an underground mine can contain dissolved carbonates. 
When hydrated lime is added in the treatment process 
to neutralize the AMD, the dissolved carbonates react 
with the lime to form insoluble calcium carbonate. 
The presence of dissolved carbonates in the untreated 
water increases both the amount of lime required to 
neutralize the AMD and the amount of sludge formed 
during the treatment process. Laboratory tests show 
that dissolved carbonates can be removed as gaseous 
carbon dioxide by aerating the AMD prior to the addi- 
tion of lime (pre-aeration) . Field tests were con- 
ducted at three AMD treatment plants. The tests 
demonstrated that pre-aeration is an effective method 
of reducing treatment costs for an AMD containing dis- 
solved carbonates. Pre-aeration has been permanently 
applied at each of the sites, and treatment costs have 
been substantially reduced, while strict compliance 
with environmental permits has been maintained. The 
chemistry of dissolved carbonates in the AMD treatment 
process, the results of the three field tests, and the 
specific modifications that were made to the treatment 
plants are discussed. 



INTRODUCTION 

Acid mine drainage (AMD) is formed by 
the weathering of pyritic materials, pre- 
dominantly FeS 2 , that are present in the 
waste found in strip mine spoil, coal ref- 
use piles, etc., as well as from exposed 
pyritic surfaces in active and abandoned 
underground mines. These pyritic materials 



!Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation Con- 
ference sponsored by the American Society 
for Surface Mining and Reclamation and the 
U.S. Department of the Interior (Bureau of 
Mines and Office of the Surface Mining 
Reclamation and Enforcement), April 17-22, 
1988, Pittsburgh, PA. 

2 T. C. Jageman is a Research Chemist, 
R. A. Yokley is a Senior Research Chemist, 
and G. W. Heunisch is a Staff Scientist, 
Research and Development Department, Con- 
solidation Coal Company, Library, PA. 



are oxidized in the presence of air (oxy- 
gen), water, and certain bacteria to yield 
large quantities of iron, sulfate, and 
acidity (Reaction 1), according to Stumm 
and Morgan (1981). 



FeS 2 (s) + 7/2 2 + H 2 - 
Fe 2+ + H 2 S0, + S0 U 2 " 



(1) 



Sulfuric acid, one product of the oxi- 
dation reaction, is readily recognizable as 
an acid. The ferrous iron produced in 
Reaction 1 is also a source of acidity as 
shown in Reactions 2 and 3 (Stumm and 
Morgan, 1981). The abiotic oxidation of 
ferrous iron (Reaction 2) proceeds slowly 
in acidic solution. In the absence of 
significant bacterial activity the acidity 
associated with ferrous iron in many deep 
mine sites is not released until the water 
is pumped to the surface for treatment. 

Fe 2+ + 1/4 2 + H + - Fe 3+ + 1/2 H 2 (2) 

Fe 3+ + 3 H 2 - Fe(0H) 3 + 3 H + (3) 



131 



The water entering a deep mine does so 
by percolating through the overburden and 
surrounding strata. If the water passes 
through limestone, it will contain alka- 
linity in the form of calcium carbonate 
and/or bicarbonate when it enters the mine 
(Parizek and Tarr, 1972). When water con- 
taining alkalinity encounters the products 
of pyrite oxidation, neutralization takes 
place (Reactions 4 and 5). 



H,S0 U + 2 CaCO, 



r 2+ , 

Ca + 



SO, 



+ Ca(HC0,) 



3 / 2 



2 + 



H 2 S0l> + Ca(HC0 3 ) 2 * Ca + 



4) 



(.5) 



SO, 



+ 2 H 2 C0 3 



This neutralization moderates the pH 
of the AMD by consuming sulfuric acid and 
may limit the solubility of ferrous iron. 
However, the products of this neutraliza- 
tion, bicarbonate ion 
are weak acids. When 
partially neutralized 
pumped to the surface 
alkaline material such 



and carbonic acid, 

AMD that has been 

in this fashion is 

and treated with an 

as hydrated lime, 



the carbonic acid and bicarbonate ion react 
with the lime to form sparingly soluble 
calcium carbonate (Reactions 6 and 7). 
Because of these reactions, excessive 
amounts of alkaline material must be used 
to achieve a pH that is favorable for 
ferrous iron oxidation and manganese 
removal . 

H 2 C0 3 + Ca(0H) 2 - CaC0 3 + + 2 H 2 (6) 

2 HC0 3 " + Ca(0H) 2 + Ca 2+ •> (7) 

2 CaC0 3 + + 2 H 2 

Work performed in our laboratory dem- 
onstrates that the AMD pumped from most 
deep mines contains dissolved carbonates 
which can be removed from solution as gas- 
eous carbon dioxide. The removal is accom- 
plished by aerating the mine water prior to 
the addition of alkaline material (pre- 
aeration) and can result in a substantial 
reduction in the amount of alkaline mate- 
rial required for neutralization. The use 
of pre-aeration to reduce AMD treatment 
costs has been reported previously (Herman 
and Korb) for a mine drainage which was 
treated to a high pH (>10) to assure mang- 
anese removal. Our work indicates that 
pre-aeration is much more broadly appli- 
cable than previously reported. 

Field Test 

An AMD treatment plant located in 
western Pennsylvania was used to investi- 
gate the utility of pre-aeration. 
site (fig. 1), AMD containing 520 
dissolved carbonate (calculated as 
pumped from a flooded deep mine. 
stream is split from the main flow 
used to form a lime slurry. The 
stream 
AMD i n 



At this 

mg/L of 

C0 2 ) is 

A side 

and is 

si urry 

is recombined with the main flow of 

an open channel which leads to a 



In the first basin an aerator mixes 
the lime slurry with the main flow of AMD 
and initiates the ferrous to ferric iron 
oxidation reaction. In the second basin, 
aerators drive the oxidation reaction 
almost to completion. 

The treated water discharges from the 
second aeration basin through an open chan- 
nel into a large settling basin. Here, 
continued oxidation takes place, and the 
insoluble ferric oxyhydroxide sludge set- 
tles out. Clarified water discharges from 
the settling basin, and the solids are 
pumped from the bottom of the settling 
basin back into the mine. 



FINE CONTROL t. - 
VALVES FOR i+ 
LIME SLURRY 



SAMPLE 

P S"rt TAP -rC>«E VALVE FOR 
GAUGE O ■ T mn FL0W 



. EXISTING LIME SLURRY CHANNEL 

-EXPERIMENTAL LIME SLURRY CHANNEL 



£ 



SuLf 




« - " DISCHARGE WEIR SETTLING BASIN 



DISCHARGE WATER 



Figure 1.— Sketch of Site 1. 



series of two aeration basins 



The field test consisted of a baseline 
test in which the plant was operated nor- 
mally. A six-day pre-aeration test was 
then performed. The AMD was treated at the 
same flow rate and same pH (8.4) as the 
baseline test. Pre-aeration was accomp- 
lished by diverting the lime slurry to the 
second aeration basin (fig. 1). In this 
configuration the first basin was used to 
pre-aerate the unlimed AMD causing dis- 
solved carbon dioxide to evolve to the 
atmosphere. Even though the residence time 
(-15 minutes) in the pre-aeration basin was 
too short for optimum removal of the 
dissolved carbonates, pre-aeration produced 
the following results: 

o 38% of the dissolved carbonates were 
removed in the pre-aeration basin. 

o Hydrated lime use was reduced 27%. 

o The concentration of unreacted ferrous 
iron leaving the aeration basin was 
reduced from 55 mg/L to 5 mg/L. 

o The sludge settling rate improved 
dramatically, and the sludge volume 
was reduced by about 50% (fig. 2). 

o The discharge water quality was main- 
tained with the concentration of total 
iron <2 mg/L and manganese <1 mg/L. 



132 



The treatment plant continued to 
operate as it was configured for the 
experiment for the next 18 months. As a 
result treatment costs were reduced by 
approximately $300,000/yr. 



1000 



900 



800 - 



700 



d 600 



500 



400 



300 



200 



100 




20 30 40 50 
TIME IN MINUTES 



60 



# 



24 HR 



Application to Other Sites 

Field tests and modifications were 
made at two additional sites. At one of 
these sites (Site 2 shown in fig. 3) the 
treatment plant was built on a hillside 
with the lime handling equipment and mix 
tanks at a higher elevation than the aera- 
tion basin, which was at the base of the 
hill. The treatment plant was designed as 
previously described (Site 1) with a small 
side stream split off to form a lime slurry 
and then recombined with the main AMD flow. 






'WtU f 

6 



ottP.lu Paw 




A, 



SETUIN& IMPOUNDMENT 




DISCHA0GE WATER 



Figure 2. --Sludge Settling Data for Site 1. 



Figure 3.— Sketch of Si I 



Recently, the treatment plant was mod- 
ified, and a permanent pre-aeration basin 
was constructed. Based on two weeks of 
operating data, chemical costs for the new 
plant are 62% less than the costs before 
pre-aeration (table 1). Approximately one- 
fourth of the savings is due to improved 
water quality in the flooded mine. One-half 
of the savings is due to preaeration. The 
remaining one-fourth is due to improved 
lime utilization. 



The AMD treated at this site has a pH 
of 4.6 and contains 200 mg/L of dissolved 
carbonates (table 2). Laboratory tests 
indicated, because of the low pH of the 
AMD, that the dissolved carbonate removal 
at this site would be rapid. Instead of 
installing a pre-aeration basin, it was 
decided to utilize the site topography. A 
pipeline was installed to transport the 
lime slurry to the aeration basin. The 
bulk of the AMD was allowed to flow over 



Table 1.— Analysis of Site 1 AMD and 
the Effect of Pre-aeration. 



AMD Sample pJi 

January 1986 

Raw 6.0 

Pre-aerated 6.4 



Fe 2 + 
(m q /L ) 



603 
585 



Total 

Dissolved 

Carbonates 

(ma CQ 2 /L) 



512 
317 



Lime 
Requi rement 
(ton/MM gal) 



8.31 
6.03 



August 1987 

Raw 6.2 447 

Pre-aerated 6.5 368 



524 
189 



5.760) 
3.19 



(*) Theoretical value calculated from the concentration 
of ferrous ion and dissolved carbonates measured in 
the raw AMD. 



133 



the hillside in a rock-lined, open channel. 
The turbulent flow of the acid water re- 
sulted in almost complete removal of the 
dissolved carbonates. 

The plant has continued to operate 
with pre-aeration for approximately 7 
months. The use of pre-aeration at this 
site reduced lime use by approximately 40% 
and resulted in a modest improvement in 
sludge settling rate (fig. 4) and sludge 
volume. Treatment costs at this site could 
be further reduced by moving the lime 
handling equipment to the base of the hill 
next to the aeration basin. This would 
allow all the AMD to be pre-aerated in the 
open channel before contacting lime. 



1000 



100 





PRE-AERATED, FIELD TEST 



20 40 

SETTLING TIME IN MINUTES 



60 



AMD HOLDING POND 



OVERFLOW CHANNEL 



SUBMERGED OUTLET PIPING (TWO) 
VALVE 




DISCHARGE PIPING 



figure 5. --Sketch of Site 3. 



The chemical reactions of these equi- 
libria provide a pathway for converting 
bicarbonate ion to carbonic acid and ulti- 
mately for converting carbonic acid to dis- 
solved carbon dioxide. Carbon dioxide has 
a very limited solubility (<1 mg/L) in an 
aqueous solution that is in equilibrium 
with the atmosphere (Cole, 1979). Aerating 
an acidic solution that contains dissolved 
carbonates causes carbon dioxide to be 
stripped from solution. This shifts the 
equilibria of Reactions 8 and 9, and more 
carbon dioxide is formed as the solution 
equilibria are re-established. In theory, 
carbon dioxide removal can continue in this 
manner until the dissolved carbonate con- 

mg/L. In practice, com- 

tne dissolved carbonates 

the 

how 



centration is <1 
plete removal of 
is unlikely, and 
AMD determines 
removal will be. 



pH of the 
difficult 



untreated 
carbonate 



Figure 4. --Sludge Settling Data for Site 2. 

At a third site, pre-aeration was 
accomplished by installing an aerator in an 
existing AMD holding pond (fig. 5). This 
pond was initially included in the treat- 
ment plant design to allow AMD, which is 
pumped to the plant from three different 
sources, to mix and equilibrate prior to 
treatment (table 3). The pond is an ideal 
retrofit a pre-aerator. Pre- 
removes approximately 60% of the 
carbonates at this site, and the 
of alkaline material used at the 
been reduced by more than 50%. 

DISCUSSION OF RESULTS 



site to 
aeration 
dissolved 
quantity 
plant has 



The chemical mechanism for the removal 
of dissolved carbonates can be explained as 
follows. Carbonic acid, one of the prod- 
ucts of the reaction of calcium carbonate 
and sulfuric acid, exists in an acidic 
solution as a component of two chemical 
equilibria (Reactions 8 and 9). 



HCO, + H" 



H 2 C0 3 



H 2 



H 2 C0 3 
C0 2 (aq) 



(8) 
(9) 



As indicated above, prior to reaching 
equilibrium with atmospheric C0 2 , the 
relative concentrations of the various 
carbonate species in AMD are governed by 
the equilibria of Reactions 8 and 9. The 
equilibrium of Reaction 9 lies far to the 
right, indicating that associated carbonic 
acid is always a minor component of the 
carbonate system (Cotton and Wilkinson, 
1972). The equilibrium of Reaction 8 
indicates that in strongly acidic solution 
(pH <4.5) dissolved carbon dioxide is the 
only significant carbonate species. At 
more moderate pH values (pH >6.4) bicar- 
bonate ion becomes the predominate species. 

The form of the carbonates in an AMD 
is critical to the design of a pre-aeration 
system because dissolved carbon dioxide is 
readily stripped from solution, but bicar- 
bonate ion is not. In order to remove bi- 
carbonate from solution the bicarbonate ion 
must react to form carbonic acid (Reaction 
10) and then dissociate to dissolved carbon 
dioxide (Reaction 11). This sequence of 
reactions takes time, particularly Reaction 
11 (Cotton and Wilkinson, 1972). 



HC0, + H" 



H 2 C0 3 



10) 



134 



Table 2.— Analysis of Site 2 AMD and the Effect of Pre-aeration 



Sample 

Pump #1 
Pump #2 

Combined AMD 
After Pre- 
aeration 



Flow Rate 
(gal /mi n) 

1,250 
1,500 



2,750 



fiM 

4.5 
4.6 



4.8 



Total 
Iron 
(m q /L ) 

206 
164 



180 



Dissolved 
Carbonates 
(ma C0,/L) 

190 
207 



Lime 

Requirement 

(ton/MM gal) 

2.7 
2.5D 



1.6 



(0 Theoretical value calculated from the concentration of ferrous 
ion and dissolved carbon dioxide in the raw AMD. 

Table 3. —Analysis of Site 3 AMD and the Effect of Pre-aeration. 



AMD Sample pJi 

Composite 

Raw 6.4 
After Pre- 6.6 
aeration 



2 + 



Fe' 

(m g /L ) 



306 
266 



Dissolved 
Carbonates 
(ma C0,/L) 



546 
233 



Lime 

Requirement 

(ton/MM aal) 



3 
1.7 



H 2 C0 3 * C0 2 (aq) 



+ HoO 



(11 



This pH dependent distribution of the 
carbonate species and the time required to 
remove the bicarbonate ion explains the 
differences in the performance of the three 
pre-aeration systems that were discussed. 
At Site 2, where AMD of pH 4.6 is treated, 
the carbonates are present as dissolved 
carbon dioxide, a rapid and nearly complete 
removal of the carbonates is accomplished 
by simply allowing the water to fall over a 
hillside. At Site 1, where AMD of pH 6.2 
is treated, approximately half of the dis- 
solved carbonates are present as bicarbon- 
ate. Agitating this water in a large basin 
with mechanical aerators removes only 65% 
of the carbonates. At Site 3 the pH of the 
water is also high (6.4), and only partial 
removal of the dissolved carbonates is 
achieved. 

An additional point worth mentioning 
is that for some mine drainages, pre- 
aeration alone can effect neutralization 
and iron removal. For mine drainages where 
the molar ratio of bicarbonate ion to 
ferrous ion exceeds 2:1, the acid produced 
by iron oxidation (Reactions 2 and 3) is 
consumed by the conversion of bicarbonate 
ion to carbon dioxide (Reactions 10 and 
11). Since this condition (high bicar- 
bonate concentration) can only occur when 
the pH of the untreated water is high (>6), 
pre-aeration can cause the iron to oxidize 
and precipitate while the pH remains at a 
level that is suitable for discharge. 

SUMMARY 

Pre-aeration has proven to be a very 
effective means of treatment cost reduction 
for deep mine AMD. In addition to the 



three sites discussed in this paper, pre- 
aeration has recently been installed at one 
other site and has been recommended for 
several others. It is anticipated that 
pre-aeration will eventually reduce AMD 
treatment costs for Consolidation Coal Com- 
pany by approximately one million dollars 
per year. 

LITERATURE CITED 

Cole, G. A., 1979, Textbook of Limnology, 
C. V. Moseley Co., St. Louis, MO, pp. 
263-271. 

Cotton, F. A. and Wilkinson, G., 1972, 
Advanced Inorganic Chemistry - A Com- 
prehensive Text, John Wiley and Sons, 
Inc., New York, NY, pp. 296-297. 

Herman, Stewart T. and Korb, Michael C, 
"The High Density Sludge Process — An 
Improved Technology to Treat Coal Mine 
Drainage," Presented at the 1983 Amer- 
ican Mining Congress Coal Convention, 
St. Louis, M0, May 1983. 

Parizek, R. and Tarr, E., 1972, Mine Drain- 
age Pollution Prevention and Abatement 
Using Hydrogeological and Geochemical 
Systems, In Proceedings, Fourth Sym- 
posium on Coal Mine Drainage Research, 
Pittsburgh, PA, pp. 72-75. 



Stumm, W. and Morgan, J. 
Chemistry, John Wiley 
New York, NY, p. 470. 



1981, Aquatic 
and Sons, Inc., 



135 






METHODS FOR DETERMINING FUNDAMENTAL CHEMICAL DIFFERENCES 
BETWEEN IRON DISULFIDES FROM DIFFERENT GEOLOGIC PROVENANCES 1 



Richard W. Hammack, 2 Ralph W. Lai, 3 and J. Rodney Diehl^ 



Abstract. --X-ray photoelectron spectroscopy (XPS), evolved 
gas analysis (EGA), and froth flotation tests were used to 
compare iron disulfides of hydrothermal, 

sedimentary/hydrothermal, and sedimentary origin. A specimen 
composed of equal amounts of pyrite and marcasite was also 
evaluated. The susceptability of iron disulfide surfaces to 
oxidation was measured using XPS and EGA techniques. XPS 
analyses indicated the following order of increasing oxidation 
rate at 21 pet oxygen and 88 pet relative humidity: 
sedimentary/hydrothermal pyrite (0.70 mg S0ij _2 /hr per gram of 
FeS2)< hydrothermal pyrite (0.83 mg S0ij _2 /hr per gram of FeS2)< 
hydrothermal pyrite/marcaslte (1.3 1 * mg S0i)_2 ,/nr P er 8 ram of 
FeS2)< sedimentary pyrite (3-53 mg S0i) _2 /hr per gram of FeS2). 
Oxidation rates measured by XPS are based solely on the 
sulfate/sulfide ratios at the surface, where oxidation is not 
inhibited by mass transfer limitations. Therefore, these rates 
are much higher than previously published rates based on bulk 
iron disulfide content. The comparison of EGA results with 
oxidation rates measured by XPS showed that for sedimentary 
pyrites, higher temperatures of SO2 evolution corresponded to 
lower oxidation rates. Weathering rates for hydrothermal iron 
disulfides appear to be independent of SO2 evolution 
temperatures. In flotation tests with an anionic 
fluorosurfactant collector, hydrothermal pyrite floated and 
sedimentary pyrite was depressed. Hydrothermal pyrite floated 
because it developed a positive surface charge in solution that 
allowed the attachment of the anionic collector. The negative 
charge developed by sedimentary pyrite in this solution repelled 
the anionic collector, depressing sedimentary pyrite. This 
research provides a better understanding of iron disulfide 
oxidation and illustrates inherent differences in physical and 
chemical properties that significantly alter the behavior of 
pyrites of different geologic provenance. 



1 Paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the 
Interior (Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), April 17-22, 
1988, Pittsburgh, PA. 



2 Richard W. Hammack is a Geologist, Pittsburgh 
Research Center, U.S. Department of the Interior, 
Bureau of Mines, Pittsburgh, PA. 

^Ralph W. Lai is a Metallurgist, and 
J. Rodney Diehl is a Physical Scientist, U.S. 
Department of Energy, Pittsburgh Energy Technology 
Center, P.O. Box 10940, Pittsburgh, PA. 



136 



INTRODUCTION 

The weathering of sulfide minerals accounts 
for most of the surface water and groundwater 
contamination that results from mining. Pyrite, 
the most prevalent sulfide mineral, is primarily 
responsible for the acid discharges from 
underground coal and metal mines, surface mine 
spoils, and tailings (refuse) disposal areas. In 
general, not all pyrites weather at the same rate. 
This fact is apparent to anyone who has compared 
unweathered pyrite from a 50-year-old metal mine 
dump with pyrite from coal refuse that becomes 
encrusted with hydrated iron sulfates and 
oxyhydroxides (pyrite weathering products) in a 
matter of days. The rate at which pyrites oxidize 
varies significantly with surface area; many 
authors (Pugh and others, 1981; Pugh and others, 
1984; and Nicholson and others, 1987) have shown 
that there is a positive, linear correlation 
between surface area and the rate of pyrite 
oxidation. Pyrites of sedimentary (low- 
temperature) origin typically display greater 
surface area, due to smaller mean particle size and 
greater surface irregularity, than pyrites of 
hydrothermal (high-temperature) origin. Therefore, 
sedimentary pyrite would be expected to be more 
reactive than hydrothermal pyrite. However, recent 
work (Esposito and others, 1987) has suggested that 
differences in surface area do not adequately 
explain differences in pyrite solubility. 

In this study, pyrites of three different 
origins (sedimentary, sedimentary /hydrothermal, and 
hydrothermal) were compared, using x-ray 
photoelectron spectroscopy (XPS) weathering tests, 
evolved gas analysis (EGA), and flotation response 
measurements. A hydrothermal specimen containing 
about equal amounts of pyrite and marcasite was 
also evaluated. The intent of this study was to 
determine if surface area differences are solely 
responsible for the wide range of observed pyrite 
reactivities, or if fundamental chemical 
differences must also be considered. We also 
wished to determine if a simple analytical 
technique, such as EGA, could be used to quickly 
predict the rate at which different pyrites 
oxidize. 



In contrast to previous studies, a surface 
analysis technique (XPS) that detects both pyritic 
sulfur and sulfate sulfur at the point of reaction 
was used. XPS provides quantitative elemental 
information with a detection limit of approximately 
1 pet of atoms comprising the outermost monolayer. 
Different formal oxidation states can be 
distinguished, based on chemical shift information. 
Chemical shifts in the sulfur (2p) electron binding 
energy (fig. 1) were used in this study to 
distinguish between pyritic sulfur (reactant) and 
sulfate sulfur (product). The XPS analysis depth 
for pyrite is estimated to be about 2.3 nm, based 
on measured inelastic mean free paths for similar 
semiconductors (Buckley and others, 1987). This 
depth corresponds to the thickness of 4.24 pyrite 
unit cells (0.54175 nm/unit cell). 




1 76 



I74 I72 



I70 I68 1 66 1 64 I62 
BINDING ENERGY, eV 

Figure 1. XPS SCAN of the sulfur (2p) region 

showing pyritic sulfur (B.E. = 163. 5eV) and 
sulfate sulfur (B.E. =169. 5eV) after 2, 6, and 
13 days weathering. 



TECHNICAL APPROACH 

XPS Weathering Tests 

The intention was to measure only oxidation on 
the surface of iron disulfide particles where mass 
transfer effects are minimized. By minimizing mass 
transfer or diffusional effects, we hoped to 
measure rates that more accurately reflect the 
chemical kinetics of iron disulfide oxidation. 
Because rates measured in this study are based on 
the amount of iron disulfide at the surface and 
"available" for reaction, these rates are 
significantly faster (10 to 20 times) than rates 
typically reported for iron disulfide oxidation. 
Previous studies have based observed oxidation 
rates on bulk pyrite content (amount of sulfate 
produced or oxygen consumed per gram of bulk pyrite 
per hour), although only a small fraction of the 
bulk pyrite was at the surface and available for 
reaction. Available pyrite can be calculated from 
the bulk pyrite content, using surface area 
estimation or gas absorption techniques, but this 
calculation involves several approximations and can 
only be regarded as an estimate. 



Most published studies (Braley, 1960; Clark, 
1965; Rogowski and Pionke, 1984; and Nicholson and 
others, 1987) have used sulfate production as a 
measure of the amount of oxidation that has taken 
place. In these studies, sulfate cannot be 
detected until it is dissolved and leached from the 
sample. This adds the complexity of sulfate 
solubility, pH-dependent sulfate adsorption, and 
sulfate transport considerations. Oxidation rates 
based on insitu detection of "available" pyrite 
and sulfate are better than rates based on "bulk" 
pyrite and leached sulfate for determining the 
mechanism and kinetics of pyrite oxidation. 
Because oxidation products were not leached from 
pyrite surfaces, the XPS technique permitted us to 
monitor, for the first time, the response of the 
oxidation rate to the accumulation of oxidation 
products. 

Iron disulfides were weathered at a constant 
humidity of 88, pet but at different oxygen partial 
pressures. The conversion of sulfide sulfur to 



137 



sulfate sulfur was monitored. For each sample, a 
percent conversion was calculated from the 
integrated photoelectron intensities for sulfate 
sulfur (binding energy = 169 eV) and pyritic sulfur 
(binding energy = 1 63 eV) : 

Percent conv.= I sulfate sulfur 

x 100 pet. d sulfate sulfur + I sulfide sulfur) 

where I = integrated photoelectron intensity. 

Because sulfate sulfur and sulfide sulfur were the 
only observed sulfur species, the equation can be 
rewritten. 



hydrothermal pyrite from the Noranda Mine, 
Quebec, Canada; and a hydrothermal marcasite from 
Joplin, Missouri. Each sample for XPS and EGA 
analysis was crushed to a -150 to +200-mesh size 
fraction and centrifuged in acetyl tetrabromide 
(2.96 specific gravity) at 1500 rpm for 20 minutes 
to separate the iron disulfide from less dense 
clays, coal, and accessory minerals. The sink 
fraction was then washed with certigrav 
(1.20 specific gravity) to remove residual bromine. 
X-ray diffraction of the cleaned iron disulfides 
indicated that some mineral contaminants remained. 
These contaminants are listed in Table 1 . 



Percent conv. = I sulfate sulfur x ^qo pet 



I total sulfur 



Evolved Gas Analysis 

Evolved gas analysis is a thermal analysis 
technique that employs a programmable tube furnace 
to heat samples in an oxidizing atmosphere. The 
concentrations of gases evolved from the sample are 
monitored with respect to sample temperature. This 
study monitored the sulfur dioxide evolved from the 
high-temperature, anhydrous oxidation of iron 
disulfides: 

2 FeS 2 + 5.5 2 — > Fe 2 3 + 4 S0 2 

Reactive iron disulfides are expected to oxidize 
and evolve sulfur dioxide at lower temperatures 
(lower activation energy) than more stable forms. 
Therefore, the sulfur dioxide evolution temperature 
can serve as a qualitative indicator of oxidation 
rate. EGA can rank iron disulfides in order of 
increasing reactivity by the anhydrous oxidation 
pathway but the correlation of sulfur dioxide 
evolution temperatures to hydrous, room temperature 
oxidation rates had not been established prior to 
this study. 

SAMPLE DESCRIPTION 

Pyrites from three different origins were 
tested in this study: a hydrothermal pyrite from 
Rico, Colorado; a pyrite of sedimentary origin that 
was later metamorphosed (sedimentary/hydrothermal 
pyrite) from the Mammoth anthracite seam in eastern 
Pennsylvania; and a sedimentary pyrite from the 
Upper Freeport coalbed, Coshocton County, Ohio. 
The hydrothermal pyrite/marcasite was from Joplin, 
Missouri. Additional iron disulfide samples that 
were not subjected to all tests included: 
sedimentary pyrites from the Pittsburgh coalbed, 
Barbour County, West Virginia (one sample) and from 
carbonaceous shales overlying the Clarion coalbed, 
Clearfield County, Pennsylvania (two samples); a 



TEST PROCEDURES 

XPS Weathering Tests 

Immediately prior to weathering, samples for 
XPS analysis were washed with boiling 4.8 N HC1 to 
remove sulfate, rinsed with methanol, and dried 
under vacuum. For each sample, thirty cylindrical 
wafers (13 mm diameter by 0.5 mm) were pressed 
under 700 kg/cm 2 pressure. Ten wafers of each 
sample were weathered simultaneously under the same 
conditions. Atmospheres used for this study 
contained 5, 10, and 21 pet oxygen (balance, 
nitrogen) at 88 pet relative humidity. Weathering 
chambers (fig. 2) were constructed so that wafers 
could be removed without disturbing or 
contaminating the atmosphere within the chamber. A 
total of three chambers were used so that samples 
could be weathered under all three atmospheres, 
simultaneously. Prior to each experiment, the 
chambers were washed with an acidified surfactant 
(sodium lauryl sulfate) solution and then rinsed 
with methanol to reduce the likelihood of 
bacterial catalysis. Periodically during the test, 
one wafer from each rod was removed from the 
chamber and placed in a Leybold-Heraeus LHS-10 
photoelectron spectrometer operated at a pressure 
of 2X1 0"^ mbar or lower. X-rays from a magnesium 
anode (MgKo = 1253-6 eV) and an analyzer pass 
energy of 100 eV were used for the acquisition of 
S(2p) and Fe(2p) spectra. Binding energy 
calibrations were carried out by adjusting the 
measured binding energy for the C(1s) spectrum of 
adventitious carbon to 284.6 eV and shifting all 
other measured binding energies correspondingly. 
Peak areas within the S(2p) region attributable to 
sulfide sulfur and sulfate sulfur were determined 
with the Leybold-Heraeus DS-5 data system. The 
percent conversion was calculated for each wafer 
and plotted versus weathering time. 



Table 1 . — Mineral contaminants in iron disulfide samples. 



Iron Disulfide 



Mineral Contaminant (s) 



Hydrothermal pyrite 

Sedimentary/ 
hydrothermal pyrite 

Sedimentary pyrite 



Calcite 

Kaolinite, calcite, quartz, 
and dolomite 

Kaolinite, quartz, calcite, 
and marcasite 



Hydrothermal pyrite/ 
marcasite 



Quartz 



138 



A^ 




Figure 2. Schematic of weathering chamber 
used in this study. 



EGA Tests 

An instrument designed specifically for 
evolved gas analysis was used. This instrument 
(fig. 3) is similar to an evolved gas instrument 
constructed by LaCount and others (1983). Major 
components include an electronic mass flow 
controller /gas blender, a programmable tube 
furnace, a quadrupole mass spectrometer, a 
programmable analog to digital (A/D) converter, and 
a microcomputer. 

The electronic mass flow controller /gas 
blender can provide selectable flow rates from 0.1 
to 200 mL/min. It can also provide two-component 
gas mixtures ranging from 0.1 to 99.9 percent. In 
this study, a 10.0 percent oxygen/90.0 percent 
nitrogen mixture was introduced into the tube 
furnace at a flow rate of 100 mL/min. 

Fifty milligrams of sample was diluted with 
3g of tungsten oxide to insure uniform heating. 
The sample was then placed in a 2.54 cm diameter by 
50 cm long quartz tube and secured with either 
glass wool or quartz wool, depending upon the 
maximum test temperature. A 32 mm (1/8 in), Type K 
thermocouple was inserted into the sample and the 
tube then placed in the furnace. Output from the 
thermocouple was conditioned by a linearizer and 
then passed to an A/D converter. 



Oxygen 




4- way valve, 

nitrogen and 

calibration gases 

Nitrogen 



5tf 



Personal 
computer 



Quadrupole mass 
spectrometer 



Figure 3. Schematic of evolved gas analysis 
instrument. 



The tube furnace used was capable of 
performing two heating ramps with selectable 
heating rates, dwell temperatures, and dwell times. 
Starting at about 70°C, the sample was heated at a 
rate of 6°C/min up to 380°C. The heating rate was 
then decreased to 3°C/min up to 720°C, where each 
run was terminated. 

Evolved gases were detected with a quadrupole 
mass spectrometer. The capillary inlet to the 
spectrometer was placed immediately downstream from 
the sample to minimize lag time between gas 
evolution and detection. The mass spectrometer was 
capable of simultaneously monitoring the ion 
current at 12 user-selected mass to charge ratios 
(M/e), although only the ion current at the M/e 
ratio of sulfur dioxide (64) is of interest here. 
This ion current was converted to partial pressure 
by multiplying by the calibration factor for sulfur 
dioxide. The partial pressure of sulfur dioxide 
and sample temperature were transmitted to a 
microcomputer where the data were converted to 
ASCII files and written to floppy disk. 
Commercially available graphics, gaussian peak 
fitting, and integration software were used to 
manipulate data. 



Table 2. — Surface oxidation rates of iron disulfides at 88 pet 
relative humidity and 5 pet, 10 pet, and 21 pet oxygen. 



Sample 


Oxidation 
(mg S0i4~ 2 g- 1 
5 pet O2 


Rate 

FeS 2 hr" 1 ) 

10 pet O2 


21 pet O2 


Sedimentary pyrite 


1 .85 




N.D. 1 


3.54 


Hydrothermal pyrite/ 
marcasite 


0.92 




1.03 


1.33 


Hydrothermal pyrite 


0.78 




0.88 


0.83 


Sedimentary/ 
hydrothermal pyrite 


0.41 




0.54 


0.70 



''Not Determined--insuff icient number of samples. 



139 




WEATHER1G TIC. HOURS 

Figure 4. XPS monitored weathering of sedimentary 
pyrite at 21 pet. oxygen and 88 pet. relative 
humidity. Solid line denotes the least squares 
regression line used for rate calculations. 




WEADERK IK HOURS 

Figure 5. XPS monitored weathering of pyrite/ 
marcasite at 21 pet. oxygen and 88 pet. 
relative humidity. Solid line denotes the 
squares regression line used for rate 
calculations. 



RESULTS AND DISCUSSION 

XPS Tests 

Plots of percent conversion of pyritic sulfur 
to sulfate sulfur versus weathering time (figs. 4- 
7) exhibited an initial linear region representing 
oxidation of about 40 to 90 percent of the 
available pyritic sulfur. Abiotic oxidation rates 
for each iron disulfide (table 2) were calculated 
from the slope of the least squares line fitted to 




200 



400 
WEATHEJflNG THE, HOURS 



Figure 6. XPS monitored weathering of hydrothermal 
pyrite at 21 pet. oxygen and 88 pet. relative 
humidity. Solid line denotes the least squares 
regression line used for rate calculations. 




200 



400 
WEATHERMGTME, HOURS 



Figure 7. XPS monitored weathering of sedimentary/ 
hydrothermal pyrite (anthracite pyrite) at 21 
pet. oxygen and 88 pet. relative humidity. 
Solid line denotes the least squares regression 
line used for rate calculations. 

the data in the initial linear region (solid line, 
figs. 4-7). Similar plots were obtained for tests 
at 5 pet and 10 pet oxygen. 

The rates in table 2 reflect abiotic surface 
oxidation as measured by XPS. Because only pyritic 
sulfur and sulfate sulfur on the surface are 
detected, this technique can be used to compare the 
oxidation rates of iron disulfides with 
significantly different surface areas. Oxidation 
rates are simply the change in the sulfate 
sulfur: pyritic sulfur ratio with respect to time. 



140 



In samples with high surface area, more pyritic 
sulfur is detected by XPS, because more pyrite is 
exposed at the surface. Therefore, the technique 
compensates for differences in surface area and 
permits the oxidation of each iron disulfide to be 
monitored under near-ideal conditions. It is 
evident from the oxidation rates in table 2 that 
sedimentary pyrite is significantly more reactive 
than pyrite of high temperature (hydrothermal) 
origin irrespective of surface area. 
Sedimentary/hydrothermal pyrite, which originally 
formed at low temperatures (< 50°C) and was later 
metamorphosed, displayed a reactivity similar to 
hydrothermal pyrite. Even the specimen containing 
roughly equal parts hydrothermal pyrite and 
hydrothermal marcasite exhibited a reactivity that 
was closer to hydrothermal pyrite than to 
sedimentary pyrite. 

The dependence of the abiotic oxidation of 
iron disulfides (initial rates) on oxygen content 
is shown in figure 8. Oxygen dependency ranges 
from no apparent dependence (Oth order) in the case 
of hydrothermal pyrite to a significant dependence 
in the case of sedimentary pyrite. Both 
pyrite/marcasite and sedimentary/hydrothermal 
pyrite exhibited a similar but slight oxygen 
dependency. The number of oxygen partial 
pressures tested in this study is insufficient for 
formulating rate equations or determining reaction 
order. However, the difference in oxygen 
dependence between pyrites of pure sedimentary 
origin and pyrites of predominantly hydrothermal 
origin appears to be significant. 



Sedimentary pyrite and pyrite/marcasite 
specimens reached an asymptotic level within the 
800-hour experiments (70 pet. conversion, fig. 4 
and 45 pet. conversion, fig. 5) where no further 
oxidation took place. The presence of an asymptote 
indicates that the mineral surface has become 
passivated. Possible explanations for this 
apparent passivation include the following: 

1 . The exhaustion of pyrite on the surface 
(applicable if XPS is detecting pyrite 
below the depth of weathering); 

2. The exhaustion of a more reactive pyrite 
form (remaining pyrite oxidizes at a slow 
rate that is undetectable within the time 
frame of these tests); 

3. In the case of sedimentary pyrite, organic 
sulfur in the form of thiophenes, thiols 
and organic sulfides would appear at the 
same binding energy as pyrite but would be 
unreactive; 

4. A metal-deficient sulfide layer forms that 
cannot be distinguished from pyritic 
sulfur, but is unreactive and passivates 
pyrite surfaces. 



OsJ 

en 



4- 



CD J 

CD 
CO 

CD 2 



en 



■ Sedimentary Pyrite 
o Pyrite/marcasite 
a Hydrothermal pyri te 
+ Sedimentary/hydro thermal 



^ 




o 



5 10 15 20 

OXYGEN CONTENT. POT. 

Figure 8. The dependence of oxidation rate on oxygen content. 



25 



141 



Explanation 1 probably is not valid because we 
have observed in depth profiles (XPS analysis with 
argon ion sputtering), weathering 10-20 nm below 
the surface of sedimentary pyrite which is at least 
four times the analysis depth for XPS. Explanation 
2 cannot be proved or disproved with current 
information. Explanation 3 may explain the 
presence of asymptotes in sedimentary pyrite plots, 
but not in hydrothermal pyrite/marcasite plots 
where no organic sulfur is present. Passivation 
due to the formation of a metal-deficient sulfide 
layer or the accumulation of oxidation products on 
the surface may best explain the asymptotes. 
Buckley and Woods (1986) found that upon 
weathering, a metal-deficient sulfide layer formed 
on pyrite which effectively passivated oxidation 
even in the presence of acetic acid. The S(2p) 
binding energy for this metal-deficient sulfide was 
1- to 1.3-eV higher than that of pyrite and 0.3 eV 
lower than bulk elemental sulfur. In the work by 
Buckley and Woods, the metal-deficient sulfide only 
formed a small shoulder on the pyritic S(2p) peak 
(7 pet of the total S(2p) intensity). The 
resolution and the signal-to-noise ratio of the XPS 
instrument used in the present study were 
inadequate for the detection of metal-deficient 
sulfide layers. 



EGA Tests 

Plots of SO2 partial pressure versus sample 
temperature are shown in figure 9. The mean 
temperature of the major peak for each iron 
disulfide increased in the following order: 
pyrite/marcasite (420°C), sedimentary pyrite 
(428°C), sedimentary/hydrothermal pyrite (470°C), 
and hydrothermal pyrite ( i )99°C). When surface 
oxidation rate (21 pet oxygen, 88 pet relative 
humidity) measured by XPS is plotted versus the 
mean temperature of the major SO2 evolution peak 
(fig. 10), two linear trends are observed. One 
trend, represented by five sedimentary pyrites, 
occurs at temperatures below 460°C with a slope of 
-0.103 mg SO4 g" 1 FeS 2 hr~ 1 /°C (r 2 = 0.991 4). The 
second trend is represented by three pyrites of 
predominantly hydrothermal origin and exhibits a 
slope of -0.007 mg SO4 g _1 FeS 2 hr~ 1 /°C (r 2 = 
0.8293). A line connecting the two marcasite- 
containing samples parallels the slope of the 
sedimentary pyrite regression line, although it 
occurs at 30oC lower temperature. Based on the 
limited data available, sedimentary and 
hydrothermal pyrite appear to represent discrete 
populations that overlap at a SO2 evolution 
temperature of 460°C. The linearity of the plot 
indicates that sedimentary pyrite specimens which 
are more susceptible to the high temperature, 
anhydrous oxidation in EGA, are also more 
susceptible to the hydrous oxidation in XPS tests 
that simulate the mine environment. However, the 
weathering rates for hydrothermal pyrite are low, 
irrespective of the major SO2 evolution temperature 
displayed by the specimen. 



UJ 

a. 
D 

CO 
CO 
UJ 

a. 
a. 



< 

Cl 
CN 
O 
CO 



Sedimentary 



Pyrite/marcasite 



Sedimentary/hydrothermal 



Hydrothermal 




T 



350 



370 



390 



410 430 450 470 

TEMPERATURE. CELSIUS 



490 



510 



530 



Figure 9. SO2 thermograms of sedimentary pyrite, hydrothermal pyrite/marcasite, hydrothermal pyrite, and 
sediment a ry/hydrdot hernial pyrite. 



142 



or 




DT 




\ 


/ 


CVJ 




CO 




LU 




LL 




I 


H 


CD 




\ 




'sr 




o 




CO 


b 


CD 




^_ 




• 


4 


LU 




t— 




CE 




CXI 






a 


-z. 




CD 




t—i 




h- 




& 


'd 


i — i 




X 




CD 






i 



a HYDROTHERMRL PYRITE 
■ SEDIMENTARY PYRITE 
a MRRCRSITE 




60Z MARCBSITE/60Z PYRITE 



360 370 390 410 430 450 470 400 510 530 550 

S02 EVOLUTION TEMPERATURE. CELSIUS 

Figure 10. Plot of SO2 evolution temperature from evolved gas analysis versus XPS monitored weathering 
rates. 



The relationship between SO2 evolution 
temperature and surface oxidation for sedimentary 
pyrites may permit EGA to be used as a quick method 
for determining surface oxidation rate. The sum of 
SO2 peak areas was plotted versus mean evolution 
temperature for 1 6 coal samples from the 
Waynesburg, Upper Freeport, and Lower Kittanning 
coalbeds (fig. 11) and for 55 overburden samples 
from Pennsylvania, West Virginia, Maryland, 
Kentucky, and Illinois (fig. 12). The sum of the 
SO2 peak areas is proportional to the total amount 
of pyrite with mean evolution temperatures within 
the 10°C interval. In both coal and overburden 
samples, most of the SO2 is evolved from pyrite in 
the 420°C and H30°C intervals, which suggests that 
the abiotic surface oxidation rate is between 
2.5- and 4.5 mg SO4" 2 g -1 FeS 2 hr~ 1 . Sulfur 
dioxide evolved at temperatures above 460°C is 
indicative of organic sulfur present in coal and 
other carbonaceous material (hydrothermal pyrite 
would also evolve SO2 in this range if present). 
When the overburden samples (fig. 12) are compared 
with the coal samples (fig. 11), the distribution 
of the SO2 peak areas appears to be 10°C higher in 
the case of the coal samples. This shift probably 
does not reflect a fundamental difference in pyrite 
reactivity, but rather a difference in sample 
preparation. The coal samples were crushed many 
months prior to EGA analysis, whereas the 
overburden samples were prepared and run within 
minutes. We have observed that with time and 
exposure to air, the SO2 evolution temperature 



I -I 
0.9- 
0.8- 
0.7 - 
0.6- 
0.5- 
0.4- 
0.3- 
02- 
0.1- 
0- 


^m 


% 


JZ/L 


\ 

i 


\ 


7 /< 


L^ 


'& 


„^ 


1 


1 



380 390 400 410 420 430 440 450 460 470 480 490 500 
TEMPERATURE CELSIUS 

Figure 11. Distribution of SO2 evolution 

temperatures for pyrite (<460°C) and organic 
sulfur (>460°C) in 25 coal samples. 



143 




370 380 390 400 410 420 430 440 450 460 470 480 490 510 520 530 
S02 EVOLUTION TEMPERATURE. CELSIUS 

Figure 12. Distribution of SO2 evolution 

temperatures for pyrite and organic sulfur in 
55 over-burden samples. 



increases. Another feature of these distributions 
is the small peak in the range of 380°C to M00°C 
that corresponds to the low temperature peak in the 
SO2 thermogram of sedimentary pyrite (figs. 9 and 
13). Pure marcasite also evolves SO2 in this range 
(fig. 13) and it is possible that the small peak 
commonly observed in the SO2 thermograms of 
sedimentary samples may represent marcasite. A 
small peak at 450°C (fig. 12) probably indicates 
that a more stable, epigenetic form of sedimentary 
pyrite is sometimes present. 

Flotation Tests 

XPS and EGA results indicate that sedimentary 
and hydrothermal iron disulfide surfaces exhibit 
significantly different reactivity. To help 
explain these differences, we used a simple froth 
flotation test to determine the charge developed on 
pyrite surfaces in aqueous solutions. This test 
consisted of placing a two-gram sample of -100 mesh 
pyrite in a 100-mL Pyrex glass flotation cell with 
a fritted bottom for gas bubbling. Each sample was 
conditioned in a solution containing 300 ppm 
anionic fluorosurfactant and 5 ppm sodium 
hydrosulf ide, and then subjected to a three minute 
flotation. The froth and the residue were 
collected, dried, weighed, and the percentage 
flotation calculated. The flotation response of a 
sedimentary and a hydrothermal pyrite is given in 
Table 3- 



LU 
C£ 
D 
C/) 
00 
Ld 
IX 
0- 



< 

0. 
CM 
O 
C/3 



Sedimentary pyrite 



Marcasite 





T 



300 



320 



340 



360 380 400 420 

TEMPERATURE. CELSIUS 



440 



460 



480 



500 



Figure 13. SO2 thermograms of hydrothermal marcasite and sedimentary pyrite. 

144 



Table 3. — Flotation response of hydrothermal and 
sedimentary pyrites. 



Sample 



Wt. % Floated 



Surface Charge 



Hydrothermal 
pyrite 

Sedimentary 
pyrite 



32 



<1 



Positive 



Negative 



Hydrothermal pyrite and sedimentary pyrite 
differed substantially in flotation response. 
Hydrothermal pyrite was readily floatable (32 pet 
floated), indicating that its surface was 
positively charged and capable of electrostatically 
attracting the anionic collector. Adsorption of 
the collector rendered the surface hydrophobic and 
allowed the flotation of hydrothermal pyrite. 
Sedimentary pyrite was depressed, indicating that 
its surface was negatively charged and not able to 
interact with the anionic fluorosurfactant. The 
surface charge must be known in order to design 
coating agents that can be applied to pyritic 
material to limit oxidation. Because the charge on 
sedimentary pyrite surfaces is negative, a cationic 
species would provide the best coating. In the 
case of hydrothermal pyrite, an anionic species 
would be a better choice. 



CONCLUSIONS 

This study showed that pyrite of low- 
temperature origin (sedimentary pyrite) differs 
from pyrite of high-temperature origin 
(hydrothermal pyrite) in many basic properties: 

1 . The observed differences in oxidation rate 
between pyrites cannot be completely 
explained by differences in surface area; 

2. The rate dependence of abiotic pyrite 
oxidation on oxygen content is 
considerably higher for sedimentary pyrite 
than for hydrothermal pyrite; 

3. In evolved gas analysis, sedimentary 
pyrite evolves SO2 at lower temperatures, 
indicating that it is more reactive than 
hydrothermal pyrite; 

4. In plots of abiotic surface oxidation rate 
versus mean SOj evolution temperature, 
sedimentary pyrites and hydrothermal 
pyrites appear to form two discrete 
populations; 



Results of this study indicate that 
passivation of pyrite surfaces occurs rapidly once 
sufficient oxidation has taken place. This 
passivation is probably caused by the formation of 
a metal-deficient sulfide layer and the accumula- 
tion of oxidation products on pyrite surfaces. 
Passivation would probably not occur under field 
conditions where 1) there is sufficient moisture 
to dissolve and transport hygroscopic iron 
sulfates, and 2) there are bacteria that can 
oxidize metal-deficient sulfide layers. 



A plot of surface oxidation rates measured by 
XPS versus SO2 evolution temperatures from EGA can 
be described by two linear trends, one trend 
represented by five sedimentary pyrites, the second 
trend represented by three iron disulfides of 
predominantly hydrothermal origin. Based on these 
linear relationships, results from quick and simple 
EGA tests can now be used to calculate surface 
oxidation rates. For example, EGA results from 
16 coal and 55 overburden samples indicate that the 
predominant surface oxidation rate for sedimentary 
pyrite is between 2.5- and 4.5 mg SOij -2 g -1 FeS2 
hr _1 . These rates are abiotic rates based on 
available pyrite (pyrite at the surface) and cannot 
be directly applied to the field without 
considering surface area and bacterial catalysis. 
However, these results can be used to augment 
existing AMD predictive techniques by providing a 
qualitative ranking of iron disulfide reactivity. 
Their origin (sedimentary or hydrothermal) is 
critical to the type of behavior exhibited by iron 
disulfides. We know from previous work that 
sedimentary pyrites display a wide range of 
reactivities. The oxidation rates displayed by the 
three hydrothermal iron disulfides in the current 
study appear to be more consistent, perhaps 
reflecting a more highly ordered crystal structure. 
It is important to recognize that the fundamental 
properties of iron disulfides, particularly of 
sedimentary pyrite, can vary widely. Iron 
disulfides are similar in gross stoichiometry, but 
little else. 



ACKNOWLEDGEMENTS 

The authors would like to thank 
Dr. Sidney Pollack of DOE's Pittsburgh Energy 
Technology Center for the X-ray diffraction 
analysis of iron disulfide specimens. We would 
also like to thank Dr. Dick Souza, mineral curator 
of the Carnegie Museum for providing 
pyrite/marcasite and marcasite specimens for this 
study. Coal samples for EGA tests were provided by 
Dr. John Renton of West Virginia University. The 
authors acknowledge the efforts of John Kleinhenz, 
Doug Zeik, Debbie Kinzler, Bob Patton, and 
Trish Steffan who made this work possible. 



Surfaces of hydrothermal pyrite are 
positively charged (floats in flotation 
tests with anionic fluor surfactant) ; 
surfaces of sedimentary pyrite are 
negatively charged (depressed in flotation 
tests) . 



145 



LITERATURE CITED 

Braley, S. A. 1960. The oxidation of pyritic 
conglomerates. Spec. Rept. to Coal Ind. 
Advisory Comm. to Ohio River Valley Water 
Sanit. Comm. Res. Proj . No. 370-6:32 p. 

Buckley, A. N. and R. Woods. 1986. The surface 
oxidation of pyrite. Applied Surface Science 
27:437-452. 

Buckley, A. N., R. Woods, and H. J. Wouterlood. 

1987. The deposition of sulfur on pyrite and 
halcopyrite from sodium sulfide solutions. To 
be published in Proceedings of the Royal 
Australian Chemical Institute Eighth National 
Convention, Sydney, Australia. 

Clark, C. S. 1965. The oxidation of coal mine 
pyrite. Ph.D. Thesis, the Johns Hopkins 
Univ. 90 p. 

Esposito, M. C, S. Chander, and F. F. Apian. 

1987. Characterization of pyrite from coal 
sources. To be published In Process 
Mineralogy VII. A. H. Vassiliou(Ed. ) . 
TMS/AIME.' 

Guilinger, T. R., R. S. Schechter, and L. W. Lake. 
1987. Kinetic study of pyrite oxidation in 
basic carbonate solutions. Ind. Eng. Chem. 
Res. 26:824-830. 

LaCount, R. B., R. R. Anderson, C. A. Helms, and 
S. Friedman. 1983. Construction and opera- 
tion of a controlled-atmosphere programmed- 
temperature reaction apparatus. U.S. Dept. of 
Energy DOE/PETC/TR-83/5. 22 p. plus appen- 
dices. 

Nicholson, R. V., R. W. Gillham, and E. J. Reardon. 
1987. Pyrite oxidation in carbonate-buffered 
solution: 1 experimental kinetics. To be 
published in Geochimica et Cosmochimica Acta. 

Pugh, C. E., L. R. Hossner, and J. B. Dixon. 1981. 

Pyrite and marcasite surface area as 

influenced by morphology and particle 

diameter. Soil Sci. Soc. Am. J. 45:979-982. 

Pugh, C. E., L. R. Hossner, and J. B. Dixon. 1984. 
Oxidation rate of iron sulfides as affected by 
surface area, morphology, oxygen concentra- 
tion, and autotrophic bacteria. Soil Science 
137:309-314. 

Rogowski, A. S. and H. B. Pionke. 1984. Hydrology 
and water quality on stripmined lands. U.S. 
Environmental Protection Agency EPA-IAG-D5- 
E763. 183 p. 



146 



INTERPRETATION OF ISOTOPIC COMPOSITIONS OF DISSOLVED SULFATES IN 

ACID MINE DRAINAGE 

Robert 0. van Everdingen and H. Roy Krouse 

1 9 

'Arctic Institute of North America, Department of Physics, 
The University of Calgary, Calgary, AB , Canada T2N 1N4 



Abstract . --Sul 


sulfide minerals in 


and waste-rock dump 


(1) 6 3l *S values sim 


sulfide minerals, a 


values similar to t 


spring waters. Exp 


waters with a wide 


that the <5 18 value 


not only on the exp 


anaerobic; submerge 


sterile or containi 


which may determine 


but also on the 6 18 


oxidation takes pla 


be interpreted as e 


oxygen in the sulfa 


molecules, whereas 


samples incorporate 


it has not been pos 


water oxygen is inc 


the actual oxidatio 


stoichiometric equa 


exchange between th 


intermediates (e.g. 


room temperature, t 


between water and s 


the observed 6 18 v 


for sulfate produce 


to widen as 6 18 va 


more negative at hi 


distances from the 


exists regarding th 


oxygen in subsurfac 


is background sulfa 


evaporites, which m 


isotope data. The 


characterized befor 


oxygen-isotope comp 


during sulfide weat 


could oxygen-iso t op 


effects of measures 


accelerate) the oxi 


deposits, and waste 



fates produced by oxidation of 


mines, mine-and-mi 11 tailings, 


s are characterized by 


ilar to those of the original 
nd (2) low or negative 6 18 


hose for sulfates in acidic 


erimental oxidation of pyrite in 


range of 6 values has shown 


of the product sulfate depends 


erimental conditions (aerobic or 


d or intermittently wet and dry; 


ng Thiobacillus f er rooxidans ) , 


the actual reaction pathways, 


value of the water in which the 


ce. The experimental results can 


vidence that 29 to 100% of the 


te was derived from water 


dissolved sulfates in field 


d 35 to 90% water oxygen. So far 


sible to determine whether the 


orporated into the sulfate during 


n process (as implied by proposed 


tions), or through oxygen-isotope 


e water and one or more 


sulfite or thiosulf a t e ) . At 


he exchange of oxygen isotopes 


ulfate ions is too slow to effect 


alues. The range in 6 18 values 


d by oxidation of sulfides tends 


lues for rain and snow become 


gher latitudes and at greater 


oceans. Some uncertainty still 


e actual 6 18 value for dissolved 


e waters. A further complication 


te from sources such as marine 


ight be detected using sulfur 


above components must be 


e proper interpretation of the 


osition of sulfate produced 


hering is possible. Only then 


e data be used to monitor the 


designed to reduce (or 


dation process in mines, tailings 


-rock dumps. 



147 



INTRODUCTION 



It h 
that oxyg 
during th 
is derive 
and in pa 
ove ra 1 1 o 
described 
equations 
(1984b) c 



2FeS +28F 

2 + 
28Fe + 



as be 
en in 
e oxi 
d in 
r t f r 
xidat 
by s 
Fo 
i te t 

e 3+ + l 



2FeS 2 +70 2 



70 2 + 
+ 2H 



en r 
co rp 
dati 
part 
on w 
ion 
ever 
r ex 
he f 

6H 

28H 

■» 2 



ealized for some time 
orated into sulfate 
on of sulfide minerals 
from molecular oxygen 
ater molecules. The 
process is usually 
al stoichiometric 
ample, Taylor et al. 
ol lowing : 

+ 30Fe 2+ +4SO^~+32H + [l] 



28Fe 3+ + 14H 2 [2] 



2+ 2- + 
Fe +4S07 + 4H 

4 



[3: 



If reaction [1] applies, all sulfate 
oxygen, Os, is derived from water 
molecules, Ow , whereas in reaction [3], 
87.5% of the sulfate oxygen is derived 
from molecular oxygen, Om , and 12.5% 
from Ow. Reaction [2], which 
regenerates Fe-*+ for reaction [1], 
produces H2O from molecular oxygen. 
However, the amount of this H2O is 
relatively minor and unlikely to affect 
the 6 18 0w and 6 18 0s values. 



It must be emphasized that the 


above equations, while convenient, may 


not describe the actual reaction 


mechanisms. For example, there is still 


considerable debate as to whether water 


dissociates and plays an oxidizing role, 


or whether exchange of oxygen isotopes 


between the water and intermediate 


sulfur species fully accounts for the 


presence of water oxygen in the product 


sulfate. According to Rimstidt et al. 


(1986), the process converting sulfide 


sulfur to sulfate may consist of as many 


as eight reaction steps, some of which 


occur at the mineral/solution interface, 


whereas others occur in the dissolved 


phase. Intermediate sulfur species may 


comprise poly sulf ides , elemental sulfur, 


polysulfonic acids, sulfite, and 


thiosulfate. Postulated pathways of 


sulfur oxidation by Thiobacilli invoke 


these intermediates (cf. review by Ralph 


1979). There is, however, the question 


whether their concentrations and/or 


lifetimes, during natural oxidation of 


metal sulfides, are sufficiently large 


for exchange reactions to alter the 


oxygen isotope composition of the final 


sulfate product. 


This paper examines published and 


new data for oxidation experiments and 


field samples, to determine how 


effectively oxygen-isotope data for 


dissolved sulfate in "acid mine 


drainage" can elucidate the conditions 


prevailing during the oxidation of metal 


sulfides. Isotopic data from laboratory 


experiments are available for oxidation 


of Na 2 S and H 2 S (Lloyd 1967, 1968); for 
oxidation of elemental sulfur (MIzutani 


and Rafter 1969); and for oxidation of 


pyrite (Schwarcz and Cortecci 1974; 


Taylor et al. 1984a; and current 



studies). Isotopic data for field 
samples have been published by Shakur 
(1979), Smejkal (1979), Taylor et al. 
(1984b), and van Everdingen et al. 
(1985); additional field samples have 
recently been analyzed by the authors. 



METHODS 



At 
abandon 
water s 
(5-cm) 
d iame t e 
stored 
glass b 
for the 
values 
values 
the sam 
collect 
and sto 
water s 
drainag 
near Co 
pond at 
(N.W.T. 
drainin 
(Britis 
Nanisiv 
Island 



the tail 
ed zinc m 
amples we 
ID piezom 
r bailer . 
in 1-L po 
o 1 1 les . 

de termin 
for disso 
for the w 
e sit e , s 
ed from d 
red in pi 
amples we 
e at two 
leman (Al 

a gold m 
) ; and f r 
g pyritic 
h Columbi 
ik lead/z 
(N.W.T. ) . 



ings 
ine 
re c 
e t er 
Th 
lyet 
The 
atio 
lved 
ater 
ampl 
rill 
asti 
re c 
aban 
bert 
ine 
om a 
exp 
a), 
ine 



depo 
in no 
ollec 
s , us 
e sam 
hy len 
samp 1 
n of 

sulf 
, res 
es of 
holes 
c bag 
ollec 
doned 
a); f 
near 
cidic 
osure 
and n 
mine 



sit of an 
rthern Onta 
ted from 2- 
ing a small 
pies were 
e and 125-m 
es were use 
& 3k S and 6 1 
at e , and 6 1 
pec t ively . 

tailings w 

and tes tpi 
s . Addit io 
ted from mi 

coal mines 
rom a taili 
Contwoyto L 

creeks 
s near McBr 
ear the 
on Baffin 



rio , 

inch 



1 
d 

8 0s 
8 0w 
At 
ere 
ts, 
nal 
ne 

ngs 
ake 

ide 



used w 
values 
+34.0° 
125-mL 
with 1 
pyrite 
dry ai 
sealed 
arms w 
After 
from o 
to det 
60 mL 
second 
de term 



n our 
ater s 
, ran 
/oo . 

vacu 
00 mL 

that 
r. T 

wi th 
ere 1 
31 da 
ne se 
ermin 
of wa 

set 
ine b 



oxidat 

with f 

ging fr 

For ea 

um flas 

of wat 

had be 

he tops 

paraf i 

eft ope 

ys , the 

t of f o 

e 6 18 0s 

ter was 

of flas 

oth 6 18 



ion exper 
our diffe 
om -33.5 
ch type o 
ks were p 
er and 10 
en finely 

of the f 
lm, while 
n to the 

water wa 
ur flasks 
; after 5 

removed 
ks and an 
Os and 6 



iments we 



rent 



;ll 



Ow 



to 

f water, two 
repared , each 
mg of 

ground in 
lasks were 

the side 
atmosphere . 
s removed 

and analyzed 
47 days , 
from the 
alyzed to 
8 0w. 



Sulfur and oxygen isotope analyses 
followed procedures described by Thode 
et al . (1961), van Everdingen et al. 
(1982), and Ueda and Krouse (1986). 

In the following sections, oxygen 
and sulfur isotope data are reported as 
per mil (°/oo) 6 18 and 6 3k S values, 
defined as: 



6 18 = 



6 3 "s 



"[ 18 0/ 16 0] 



sample 



[ 18 0/ 16 0] 



[ 34 S/ 32 S] 



- 1 



standard 



sample 



[ 34 S/ 32 S] 



standard 



x 10 3 



x 10 3 



;34, 



The usual standards for 6 aH S and 6 10 
are troilite from the Canon Diablo 
meteorite (CDT) , and water approximating 
the mean isotopic composition of the 
ocean (SMOW). 



148 



Table 1. --Isotope data from oxidation experiments 



SOURCE 



DURATION 
(days) 



6 18 0s 
7oo SMOW 



6 18 0w 



/oo 


SMOW 


-33 


.5 


-19 


.5 


+ 17 


.0 


+ 34 


.0 


-26 


. 2 * 


-13 


. 1 * 


+ 22 


3 * 


+40 


5 * 


-10. 


9 


-10. 


9 


-10. 


9 


-10. 


9 


-10. 


4 


-10. 


9 


-10. 


4 


-10. 


9 


-10. 


4 


-7. 





-7. 





-0. 


8 


-0. 


8 


-0. 


8 


-8. 





+ 127. 





-4. 


7 


+ 20. 





+ 34. 






1 / 
/oo 



Current study - PYRITE 
(submerged/aerobic) 



Taylor et al., 1984a - PYRITE 

(submerged/anaer . /sterile/+Fe _) 
(submerged/aerobic/sterile/+Fe ) 
(submerged /aerobic /sterile) 
(submerged/aerobic/+T.ferro.) 
(submerged/aerobic/+T.ferro.) 
(wet/dry, sterile) 
(wet/dry, sterile) 
(wet/dry, + T.ferro.) 
(wet/dry, + T.ferro.) 

Mizutani & Rafter, 1969 - SULFUR 
(in aerated water with soil) 



Schwarcz & Cortecci, 1974 
(in aerated water) 



Lloyd, 1967 - H S 

(through aerated water) 



PYRITE 



31 

31 

31 

31 

547 

547 

547 

547 



30 
26 
27 
27 
27 
29 
29 
29 
29 



14 
17 
15 
18 
21 



14 
14 



-5 
+0 
+ 16 
+ 22 
-14 
-5 
+ 15. 
+ 29. 



+ 3, 
+ 71. 

+ 2. 
+ 18. 
+ 2 9, 



-6.9 
-6.0 
-4.8 
+ 1.8 
+0.5 
+ 1, 
-0. 
+6. 
+ 7, 



-7.6 
-7.0 
-1.3 
-0.4 
+ 2.0 



+ 27 


.6 


+ 20 


.2 


-1 


.0 


-11 


.4 


+ 12 


.0 


+ 7 


.4 


-6 


.8 


-11 





+4 





+4 


9 


+ 6 


1 


+ 12. 


7 


+ 10. 


9 


+ 12. 


2 


+ 9. 


5 


+ 17. 


5 


+ 18. 


1 


-0. 


6 


0. 





-0. 


5 


+0. 


4 


+ 2. 


8 


+ 11. 


5 


-56. 





+ 6. 


7 


-2. 





-4. 


7 



Values at end of experiment reflect evaporation effects 



+ 40- 



O 

o 

vJ5 



V> 

O 

CO 
60 



H-127,+71) 



+ 20- 




8 18 w (%o SMOW) 



,2- 



Figure l.--6 18 0s values of SO^ - versus 
<5 18 0w value! 
experiments 



6 18 0w values of H 2 for oxidation 



DATA FROM LABORATORY EXPERIMENTS 

Isotopic data from oxidation 
experiments are listed in table 1 and 
illustrated by figure 1. 

Previous Work 

Lloyd (1967, 1968) concluded that 
the ratio of water oxygen to molecular 
oxygen used in the oxidation of Na 2 S , at 
25°C in water under an oxygen 
atmosphere, was 1:2. He suggested that 
the initial reaction (not balanced) was 



+ H 2 + 2 



SO 



2- 



[4] 



The isotope fractionation [em] for the 
incorporation of molecular oxygen was 
deduced to be -8.7 /oo. Lloyd also 
found that the subsequent reaction 



2- 
SO^ + 



JO, 



SO 



2- 



[5] 



149 



slow, and that sulfite 
sotope exchange with 
one week at room 
uch as 10 times 
change between sulfate 
nal <5 18 0s for the 

depend on t he 

( 1 ) exchange of 
tween sulfite and 
dation of sulfite to 
d be noted here that 
ion of S0^~ are pH 



Oxidation of HiS in aerated 
waters with three different 6 lo 0w 
values resulted in 64 to 76 percent 
of the sulfate oxygen being derived 
from Ow (Lloyd 1967). The isotope 
fractionation [ew] for incorpora- 



was comparatively 


unde rwen t 


oxy gen-i 


the wa t er 


(80% in 


temperature), as m 


faster than the ex 


and wa ter . 


The fi 


sulfate would thus 


c cmpe ting 


rates of 


oxygen isotopes be 


water , and 


(2) oxi 


sulfate . 


I t shoul 


bui ldup an 


d re tent 


dependent . 





the A value was smallest (1 /oo) for the 



experiment using water with a 6 
value of +17 /oo . 



Ow 



tion of water oxygen into S0J 



was found 



to be zero. The A values ( 6 18 0's minus 
6 18 0w) for individual samples ranged 
from -4.7 to +6.7°/oo (table 1). 

Eacterially mediated oxidation of 
elemental sulfur in an aerated 
soil/water slurry at 30°C produced 
sulfate with 6 18 0s values almost 
identical to 5 18 Ow (Mizutani and 
Rafter 1969). Sulfur isotope 
fractionations between sulfur and 
sulfate were less than 2.3 /oo; A 
values ranged from -0.6 to +2.8 /oo 
(table 1). 



Oxidation of finely ground 


pyrite at 25°C in aerated slurries 


made with waters with two different 


6 18 0w values revealed that & Os 


increased about 0.6 /oo for each /oo 


increase in 6 18 0w (Schwarcz and Cortecci 


1974). The A values ranged from -56.0 


to +11.5 /oo, and approximately half the 


sulfate oxygen was deduced to have been 


derived from the water (table 1). 


A series of experiments was 


designed by Taylor et al. (1984b) to 


elucidate the pathways of pyrite 


oxidation. They used both aerobic 


and anaerobic sterile conditions, as 


well as aerobic conditions with 


Thiobacillus ferrooxidans. The pyrite 


was either submerged or subjected to the 


alternating wet and dry conditions often 


found in mine environments. For 


individual experiments, the A values 


ranged from +4 /oo (anaerobic, sterile), 


to +18°/oo (aerobic, with T.ferro.). 


The isotope fractionation (cm) for 


incorporation of molecular oxygen into 


sulfate was found to be -4.3 /oo for 


chemical (abiological ) oxidation, and 


-11.4 /oo for bac t e r ially-media t ed 


oxidation (for an "average" value of 


-7.9°/oo) . 



This Study 

In our experiments, the samples 
extracted after 31 days showed A values 
ranging from -11.4 to +27.6 /oo; for the 
samples extracted after 547 days, the A 
values ranged from -11.0 to +12.0 /oo; 



The A values 


+11.1 /oo for sam 


(Smejkal 1979); f 


for samples from 


(Taylor et al. 19 


+23.0 /oo for sam 


Canada (van Everd 


from +4.0 to +22. 


analyzed during o 


values for sample 


area (N.W.T.) , an 


(1979) , were all 


+26.4 to +34.0°/o 



DATA FOR FIELD SAMPLES 

Isotopic data for most of the field 
samples are listed in tables 2 and 3, 
and illustrated in figure 2. 



ranged from +3.4 to 
pies from Bohemia 
rom +2.6 to +15.2°/oo 
Colorado and California 
84b) ; from +10.2 to 
pies from W. and N. 
ingen et al. 1985) ; and 
/oo for the samples 
ur current study. The A 
s from the Pine Point 
alyzed by Shakur 
higher, ranging from 
o . 



DISCUSSION 

Isotope Balance Calculations 

To help determine the relative 
fractions of sulfate oxygen produced by 
the model reactions [1] and [3] during 
the various experiments, the isotopic 
composition of the oxygen incorporated 
in sulfate during sulfide oxidation can 
be expressed by the isotopic balance 
equation : 

(5 18 0s = Y * (6 18 0w + ew) + (1-Y) * 

* (0.875 * ( 6 18 0m + em) + 

+ 0. 125 * (6 18 Ow + ew)) [6] 



o 

if) 

o 



O 
oo 



+20 


A 








**-- 


x\ 


\r' 







,-r* — "p* Sx 


+ 










v'/ 






50^^lT 


/ 


v BOHEMIA 
D CALIF. /COLO. 




^T 6* y 




| WEST SHASTA 




P&r 




+ SOUTH BAY 








® COLEMAN 
Px PAINT POTS 




/n* / 




B CUSH 


-20 






Ex ENGINEER CK. 






a PINE POINT 








Gx GOLDEN DEP. 








Lx LUPIN 




/100% 

/ 1 


i 


NxNANISIVIK 



-20 



-+•20 



8 18 w (%o SMOW) 



Figure 2.--6 18 0s values of S0i,"versus 
6 18 0w values of 1^0 for field samples. 



150 



Table 2.-- Isotope data for field samples 



SOURCE 





Sulfate 


Water 


6 34 S 


6 18 


6 l8 


/oo CDT 


°/oo SMOW 


°/oo SMOW 



1 / 
/oo 



A - ACIDIC SPRINGS 



PAINT POTS, B.C. 

Spring #12, 3-7-81 

Spring #12, 28-4-84 
GOLDEN DEPOSIT, N.W.T. 

Spring Water 
ENGINEER CREEK, YUKON 

Acid Discharge 
CUSH PROPERTY, B.C. 

Water 80158 

Water 80159 

Water 80160 
NANISIVIK, N.W.T. 

Acid Creek #2 



+ 10.8 
+ 9.4 

-22.8 

-23.2 

+ 1.3 
+0.9 
+ 1. 1 

+ 15.3 



-9.4 
-9.5 

+ 0.5 

+ 0.5 

-6.2 
-9.4 
-7.2 

-16.4 



20.5 


+ 11.1 


19. 7 


+ 10.2 


17.7 


+ 18.2 


22.5 


+ 23.0 


21.0 


+ 14.8 


20.4 


+ 11.0 


20.5 


+ 13.3 



!0.4 



+ 4.0 



B - MINE WATERS 



W. and N. BOHEMIA (Smejkal (1979) 
Michael coal mine 
Antonin coal seam 
Josef coal seam 

Hromnice quarry, pyritic slates 
Berk quarry, pyritic slates 
Tisova sulfide mine 
Svornost sulfide mine 
Jachymov, Geier sulfide vein 

COLORADO/CALIFORNIA (Taylor et al. 1984a, Fig. 2) 
Mine waters Minima 
Maxima 



COLEMAN, Alberta 
CC Mine Drainage 
WCC Mine Drainage 



+ 1.6 
+0.8 
+0.1 
-6. 1 
-2. 1 
-4. 1 
-4.7 
-5.9 



12.3 
-2.7 



W. SHASTA, California (Taylor et al. 1984a, Fig. 2) 

Mine waters Minima -7.6 

Maxima +5 . 7 



+ 19.6 
+ 18.6 



+ 1.9 
+3.0 



-9.5 


+ 11.1 


-9.5 


+ 10.3 


-9.5 


+ 9.6 


-9.5 


+ 3.4 


-9.5 


+ 7.4 


-9.5 


+5.4 


-9.5 


+ 4.8 


-9.5 


+ 3.6 


-15.7 


+3.4 


-15.1 


+ 12.9 


-11.1 


+ 2.6 


-9.8 


+ 15.2 


-19. 1 


+ 21.0 


-19.0 


+ 22.0 



LUPIN MINE, N.W.T. 
Tailings Pond 

PINE POINT, N.W.T. (Shakur, 1979) 
Water from open pit 



+ 3.2 



+ 14 
+ 15 
+ 16 
+ 11 
+ 16 
+ 17 
+ 18 
+ 12 
+ 11 
+ 16 
+ 15 
+ 17 
+ 18 
+ 17 



■6.3 



+ 9.8 

+ 8.6 

+ 9.8 

+5.2 

+ 12.0 

+ 11.0 

+ 11.0 

+8.2 

+ 4 

+ 11 

+ 8 

+ 10 



17.6 



+ 11.3 



4 
4 

8 
4 

+ 12.0 
+ 11.0 



-22.0 


+ 31.8 


(ave . ) 


+30.6 




+ 31 .8 




+ 27.2 




+ 34.0 




+ 33.4 




+ 33.0 




+ 30.2 




+ 26.4 




+ 33.4 




+ 30.8 




+ 32.4 




+ 34.0 




+ 33.0 



151 



Table 3.-- Isotope data for South Bay mine tailings 



WATER SAMPLES 



Piezometer # 



Sulfate 



Dep 


el 


, m 


1 .3 


_ 


2 


2 


1 .3 


- 


2 


2 


1 .4 


- 


2 


3 


1 .8 


- 


2 


7 


1 .2 


- 


2 


1 


1 .2 


- 


2 


1 


4.4 


- 


5 


1 


10.3 


- 


10 


6 


4.8 


- 


5 


1 


3.8 


- 


4 


6 


9.8 


- 


10 


1 


7. 9 


- 


8 


2 


5.5 


- 


5 


8 


5.9 


- 


6 


8 


7.0 


- 


7 


9 


7.3 


- 


8 


1 


4.4 


- 


5 


2 


16.9 


- 


17 


2 


9.9 


- 


10 


2 


5.4 


- 


6 


2 


4. 1 


- 


4 


9 


8.9 


- 


9 


2 


5.9 


- 


6 


2 


10.4 


- 


11 





2.6 


- 


2 


9 



6 3 "S 
7oo CDT 



6 18 
7oo SMOW 



Water 




6 18 


A 


/oo SMOW 


/oo 


-11.7 


+ 10.8 


-11.5 


+8.2 


-12.9 


+ 9.6 


-12.8 


+ 6.5 


-15.5 


+ 9.8 


-14.9 * 


+ 7.3 


-12.4 


+ 10.9 


-12. 1 


+ 14.5 


-12.5 


+ 6.8 


-12.0 


+ 8. 1 


-10.8 * 


+ 7.0 


-12.3 


+ 7.4 


-14.9 


+ 10. 1 


-12. 7 * 


+ 9.7 


-14.2 


+ 15.1 


-14.8 


+ 13.8 


-14.2 


+ 11.4 


-14.8 


+ 18.0 


-13.3 


+ 9.5 


-11 .4 


+ 9.9 


-12.7 


+ 15.0 


-14.0 


+ 11.4 


-13.4 * 


+ 10.2 


-12.3 


+ 9.6 


-14.7 


+ 16.7 



H 


- 


5 


H 


- 


5 


H 


- 


6 


H 


- 


7 


H 


- 


8 


H 


- 


8 


M 


- 


4 


M 


- 


5 


M 


- 


5b 


M 


- 


7a 


M 


- 


7b 


M 


- 


8 


M 


- 


9 


M 


- 


11 


M 


- 


18 


M 


- 


24a 


M 


- 


24b 


M 


- 


2 


M 


- 


32 


M 


- 


40 


M 


- 


41 


M 


- 


43 


M 


- 


45 


M 


- 


4 


M 


- 


49 



+ 4.5 
+4.6 
-0.8 
+0.2 
-0.6 
-0. 1 
+0.6 
+0. 1 
-0.9 

0.0 
+0.6 
-0. 1 
-0.2 
+0.9 
+8.3 
-0.2 

0.0 
+0.3 
+ 0.1 
+ 2.5 
+ 0.6 
-0.2 
-0. 1 
-0.3 
-0.6 



-0, 
-3, 
-3, 
-6. 
-5, 
-7. 
-1, 
+ 2. 
-5, 
-3, 



-3.8 
-4.8 



-4.7 

-3. 1 

+0.8 

-1.0 

-2 

+ 3 

-3 

-1 

+ 2 

-2 

-3 

-2.6 

+ 2.0 



TAILINCS SAMPLES 
Source 



2 Tl 
2 T2 
4 T 

26 T 

27 T 
40 T 
43 Tl 
43 T2 
46 Tl 
46 T2 
46 T3 





Dep 


th, 


m 





41 


_ 





45 


1 


.83 


- 


2 


.43 


1 


.52 


- 


2 


13 


1 


.52 


- 


2 


.13 


1 


.52 


- 


2 


13 


1 


.52 


- 


2 


.13 





.43 


- 





50 


1 


.22 


- 


2 


.43 





41 


- 





50 





.51 


- 





60 


1 


52 


- 


2 


13 



State 
of 
oxidat ion 

OX 
UNOX 
UNOX 
UNOX 
UNOX 
UNOX 

OX 
UNOX 

OX 

OX 
UNOX 



Sulfide 

6 34 S 
°/oo CDT 



+0.8 
+0.8 
+ 1. 1 
+0.5 
+ 1.2 
+0.9 
+ 1.2 
+0.8 
+ 1.2 
+ 1.0 
+ 1.3 



Sulfate 

6^S 6 18 



/oo CDT 

+ 1.4 
+ 0.4 
+0.3 
-4.8 * 
+0.6 
+0.6 
+ 1.8 



+ 0, 
+ 1, 



°/oo SMOW 

-4.8 
+ 0.8 
-0.9 
+4.0 
-2.6 



+ 1.0 
+0.7 



-5 
+4 
+ 5 
-4 
-6 



CONCENTRATE SAMPLE 



OX 



+ 0.4 



-1.3 



+ 0.1 



* - Average of two determinations 

in which Y is the fraction of sulfate 
ions produced by reaction [1]; em and ew 
represent the shifts in 6 18 0m and 6 18 Ow 
during incorporation of Om and Ow into 
sulfate . 

Using em = -8.7 /oo and 
ew = 0.0°/oo (Lloyd 1967), and the 
6 18 0m = 5 18 of atmospheric 
2 = +23.8°/oo (Horibe et al. 1973), 
equation [6] can be rewritten as 

6 18 0s = 6 18 0w * (0.875 * Y + 0.125) + 



+ (1-Y) * 13.21 



[7] 



Lines for 100, 75, 50, 25, and percent 
contribution by reaction [1], calculated 
using equation [7], have been plotted in 
figures 1 and 2. These lines converge 
and intersect at the point representing 



6 18 0s = 6 18 0w « +15°/oo. The 
convergence suggests that at low 
latitudes, where S 18 Ow values are closer 
to /oo, it will be more difficult to 
distinguish contributions from 
individual reactions than at high 
latitudes, where 6 18 0w values may 
approach -30 /oo. It is noted that in 
the earlier analysis of van Everdingen 
and Krouse (1985), ew was chosen as 
+2.6 /oo, as suggested by data of Taylor 
et al. (1984a). However, with that ew 
value, some of our recent data would 
fall below the 100 percent reaction [1] 
line on a 6 18 0s vs. 6 18 0w plot. 
Therefore, an ew value of zero seems 
more appropriate. 

Equation [7] can be rewritten to 
allow calculation of Y from the 6 18 0s 
and 6 18 0w values for individual samples: 



152 



Y = (6 18 0s - 0.125 * 6 18 0w - 13.21)/ 

/ (0.875 * <5 18 Ow - 13.21) [8] 

Finally, the fraction (X) of sulfate 
oxygen derived from Ow can be calculated 
either from 



X = 0.875 * Y + 0. 125 



[9] 



or directly from the isotopic data for 
Os , Ow, and Om : 



X = (6 18 Os - ( 6 18 0m + cm)) / 
/ (6 18 0w - ( 6 18 0m + em)) 



[10] 



pom 

reac 

2 in 

sulf 

and 

data 

100% 

reac 

to 1 

deri 

[9]) 

data 

Poin 

88% 

reac 

sulf 



The p 
ts in 
tion- [ 
dica te 
ate pr 
[3]. 

(fig. 

of th 

tion [ 

00% of 

ved f r 

The 

(fig. 
t Poin 
of the 
tion [ 
ate ox 



osit ion of 
relation t 
1 ] lines i 
s the prop 
oduced by 
The spread 

1) shows 
e sulfate 
1]. This 

the sulfa 
om Ow (cal 

dis tr ibut 

2.) , with 
t se t , ind 

sulfate w 
1], with 3 
ygen provi 



individual data 


o the percent- 


n figures 1 and 


ortions of the 


reactions [1] 


of the experimental 


that between 19 and 


can be produced by 


corresponds with 29 


te oxygen being 


culated by equation 


ion of the field 


the exception of the 


icates that 26 to 


as produced by 


5 to 90% of the 


ded by Ow. 



Complications in Field Situations 



Sulfate 


in waters associated with 


sulfide mineralization may have been 


derived from 


several sources (fig. 3). 


In the case 


of carbonate -hos ted 


metal-sulf id 


e deposits, anhydrite (or 


gypsum) may 


have been dissolved earlier 


from associated marine evaporite strata 


Under anaero 


bic conditions in the 


presence of 


a suitable carbon source, 


sulfate reduction may occur, enriching 


the remainin 


g S0£ in both 31+ S and 18 
1987). Under more aerobic 


(cf . Krouse 


condi t ions , 


sulfur in lower valence 


states can oxidize, incorporating 


oxygen atoms 


from dissolved Oy and 


from H 2 . 



SULPHATE 
MINERALS 



n 



HS" 



METALS 



Small S \ 
Isotope 1 
VFractionation/ 



METAL 
SULPHIDES 



DISSOLUTION PRECIPITATION 
/ Small S and \ 
\ Isotope Fractionation/ 



REDUCTION 
Large S and \ 
Isotope 1 
Fractionation / 



TT? 

OXIDATION 
I From \ 
ImOandoJ 



S° 

ORG S 

SO. 



T 



so* 



Figure 3. — Interactions between natural 
subsurface sulfate sources. 

Interpreting isotopic data solely 
on the basis of me tal-sulf ide oxidation 
ignores the possibility that the water 
in which the oxidation process takes 
place may contain dissolved sulfate from 



non-sul 
suspe c t 
inf orma 
and the 
sulfate 
Th 
is illu 
samples 
carbona 
Point ( 
values 
respect 
ar ea . 
with [S 
local s 
anhydr i 
sulfide 



fide sources. Where this is 
ed, it is necessary to obtain 
tion on both the concentration 
6 18 Os for the "original" 



e compl 
s trated 
from a 
te-host 
Shakur 
varied 
ively ( 
The <5 -v 
0|~] an 
ources 



te, and 
s , both 
+20 u /oo (Sasak 



exity 
by t 
pump 

ed Pb 

by 6 

fig. 
alues 
d [CI 
of SO 
oxid 
with 
i and 



f natural systems 




data for water 




open pit in the 




n deposit at Pine 




The 6 3tf S and 6 18 


s 


o and 8 /oo, 




, over a small 




ended to increase 




The two known 




are dissolution o 


f 


ion of metal 




3 S value s near 




rouse 1969) . 





+ 20 



+ 10 



o 






EVAPORITES 







&180 w 

-22%o 



I SULPHIDE 
ORE 



•• 



.<£ 



O 



S 34 S HS" 
-20%o 



O 



± 



-10 



+10 

8 18 0c (%o) 



Figure 4. — 6 34 S versus 6 18 0s for So£" 
samples from an open pit near Pine 
Point, N.W.T. (after Shakur 1979). 
Ranges of 6 31 *S values for Pb-Zn ores, 
and of 6 31 * S and 6 18 values for 
evaporite sulfate are indicated. 



A graphic tool for identif 


ying 


sulfate sources is the plotting 


of 


6-values versus the inverse of 


[S0|-] or 
ear 


[CI ] concentrations. If a lin 


behaviour is invoked, the Y-intercept 


corresponds to the 6 -value for 


one of 


the sources (cf. Krouse 1980). 


For the 


Pine Point 6 18 0s data, figure 5 


shows 


that the 6 18 Os value at the intercept is 


2 to 3 per mil higher than expected for 


the evaporite sulfate (Sasaki and Krouse 


1969). Note that the line comp 


u ted with 


the points for the two lowest sulfate 


concentrations excluded, gives 


an 


intercept close to the evaporit 


e value . 


It can also be argued that 


the 


6 18 Os values for many of these 


samples 


were somewhat elevated by bacterial 


sulfate reduction, because the 


Eh values 


of these waters average -250 mV 


(Shakur 


1979). This would tend to give 


a lower 


Y-intercept, since higher 6 18 


shifts 



153 




Other Uncertainties 



0.003 



Figure 5--.S Os versus the inverse 
of sulfate concentration, for water 
samples from an open pit near Pine 
Point, N.W.T. (after Shakur 1979). 
The dashed regression line 
represents the closed circles only 
(see text for discussion). 



woul 


cone 


in f 


gene 


sulf 


the 


mini 


wher 


6 3 "S 


that 


6 18 


in f 


valu 


tren 


by t 


Ther 


oxid 


vale 


has 


Anal 


dist 



sign 

6 3 "S 

bact 

al. 

to b 

sulf 

appr 

sulf 

of s 

be i 

and 



d be associat 
entration. T 
igure 5 imply 
rated by oxid 
ur. The sulf 
source for th 
mum 6 31 * S valu 
eas figure 4 
values were 
the two samp 
s values and 
igure 5 also 
es , and that 
d that differ 
he rest of th 
e f ore , one ma 
a t ion o f ano t 
nee sulfur, w 
produced some 
ysis of water 
ance from the 
ificant conce 
-depleted HS 
erial sulfate 
1979) . Its r 
e a good cand 
ate source . 
eciated that 
ate con t r ibut 
ulfide ores i 
mposs ible wit 
iso topic data 



ed with 
he lowe 

that s 
at ion o 
ide min 
is , bee 
e is ab 
shows t 
found f 
les wit 
sulfate 
have th 
they s e 
s from 
e data 
y concl 
her sou 
ith a 1 

of the 
s from 

mining 
nt rat io 
, gener 

reduc t 
e-oxida 
idate f 
It can 
quant if 
ion fro 
n such 
hout de 



lower sulfate 
r 6 18 0s values 
ome sulfate was 
f lower valence 
erals cannot be 
ause their 
out +17 /oo, 
hat lower 



or SO 



Note 



h the lowest 
concentrations 



e lowest & 



34 



em to fit a 
that suggested 
in figure 4. 
ude that 
rce of lower 
ower 6 3 S value, 

sulfate . 
boreholes some 

area revealed 
ns of 
a ted by 
ion (Weyer et 
tion would seem 
or the third 
be readily 
ication of the 
m the oxidation 
a system would 
tailed chemical 




2 and 
are un 
of mar 
Those 
Colema 
their 
and ev 
their 
occur 
inc lud 
locate 
Nanisi 
the lo 



f ou 
3, t 
doub 
ine 
for 
n mi 



ill 



apor 

vici 

near 

e th 

d on 

vik 

west 



r field 
hose fo 
t edly a 
sulfate 
the Gol 
nes are 
s value 
ites ar 
nity . 

the re 
e South 

the Ca 
and Pai 

6 18 0s 



dat 
r th 
ffec 
, as 
den 

pos 
s ar 
e kn 
No m 
main 

Bay 
nad i 
nt P 
valu 



a list 
e Pine 
ted by 

discu 
Depos i 
sibly 
e rela 
own to 
ar ine 
ing si 

and L 
an Shi 
ot s di 
es in 



ed i 

Poi 

the 

ssed 

t an 

af fe 

tive 

occ 

sulf 

tes ; 

upin 

eld, 

scha 

tabl 



n tables 

nt area 
presence 
above . 

d the 

cted ; 

ly high, 

ur in 

a tes 
these 
mine s , 
and the 

rges with 

e 2. 



rcgar 
e qua t 
molec 
sulf i 
exper 
atmos 
1973) 
6 18 0m 
usual 
be dr 
kine t 
react 
Deter 
water 
diff i 
sampl 
int ro 
the, re 
whe th 
repre 
forma 
shoul 
as it 
inter 
plots 



Some un 
ding th 
ions [6 
ular ox 
de oxid 
iment s , 
pher ic 

has be 

value 
ly not 
as t ical 
ic iso t 
ions , o 
mining 
s pr ese 
cul t ies 
e can b 
ducing 

is sti 
er the 
sent s t 
tion . 
d be ma 

bears 
pretati 

for fi 



cert 
e 6 U 

1 an 
ygen 
atio 

the 

2 ( 
en a 
of d 
de t e 
ly a 
ope 
r by 

6 18 
nts 

dur 
e ob 
atmo 
11 t 
meas 
hat 
Neve 
de t 
dire 
on o 
eld 



ain ty 
' Om va 
d [10] 

parti 
n . In 

6 18 0m 
+23.8° 
s sumed 
issolv 
rmined 
f f ec te 
effect 

excha 
m valu 
a chal 
ing sa 
tained 
spher i 
he que 
ur ed 
exis t i 
r thele 
o meas 
ctly o 
f 6 i8 
sample 



stil 
lue 

for 
cipa 

lab 

val 
/oo , 
H 
ed o 
, al 
d ei 
s du 
nge 
es i 
leng 
mpli 

wit 

c CO 

stio 




II 



ng d 
ss , 
ure 
n th 
s ve 
s . 



1 exis 
to be 

disso 
ting i 
orator 
ue for 

Horib 
owever 
xygen 
though 
ther b 
ring o 
proces 
n natu 
e beca 
ng. E 
hout 
ntamin 
n as t 
m valu 
uring 
at temp 
this p 
e 
rsus 6 



ts 

used in 
lved 

n 

y 

e et al . 
, the 
is 
it may 

y 

xidation 
ses . 
ral 

use of 
ven if a 

at ion , 

o 

e 

sulfate 

ts 

arameter 



Ow 



the 

the 

prop 

wate 

inco 

also 

shal 

to s 

some 

(ref 

inf i 

comp 

disc 

subs 

comp 



"Wil 

incl 

oxyg 

rati 

Lloy 

bell 

have 

ques 

othe 

deta 

vari 

r eso 

outl 

expe 

Os d 

oxid 

e lem 

sulf 

auth 

of c 

expe 

(in 

anal 

inte 

form 

bef o 

oxid 

with 



In 
wate 
t ime 
erly 
r th 
rpor 

be 
low 
how 

sea 
lect 
ltra 
ar ed 
harg 
ur f a 
lete 

Hoi 
1 fi 
ud in 
en o 
o of 
d's 
eve 

pro 
tion 
r qu 
iled 
ety 
lve 
ined 
rime 
er iv 
atio 
enta 
ur s 
or s 
olle 
rime 
suff 
ysis 
rmed 
ed d 
re t 
ize 

the 



addition it could be argued that 
r associated with the sulfate at 
of sample collection may not 
reflect the 6 18 0w value of the 
at provided the oxygen for 
ation into the sulfate. It should 
noted that water samples from a 
tailings deposit can be expected 
a relatively wide range of, and 
sonal variation in, 6 I8 0w values 
ing 6 18 Ow variations in 
ting rain and snowmelt), as 

to water from natural acidic 
e or mine drainage, in which 
ce mixing would be more 



ser 
eld 
g is 
f th 

1 :2 
expe 
that 
vide 
, ev 
es t i 

ana 
of f 
some 

abo 
ntal 
ed f 
n of 
1 su 
peci 
in c 
ague 
nts 
icie 
), o 
iate 
urin 
hey 
f urt 

sur 



e t a 
stud 
o t op 
e ai 

of 
rime 

the 
d an 
en w 
ons . 
lysi 
ield 

of 
ve . 

dat 
rom 

dif 
lfur 
es i 
olla 
s . 

shou 
nt q 
ne o 

sul 
g th 
have 
her 
roun 



1. (1 
ies o 
e mea 
r and 
those 
nts?" 

data 

answ 
hile 

Car 
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the u 

The 
a on 
0m an 
f eren 
, and 
s be i 
borat 
Futur 
Id be 
uant i 
r nor 
fur s 
e oxi 

a ch 
or to 
ding 



979, p 
f pyri 
sureme 

water 

parti 
The 

prese 
er to 
raisin 
ef ul s 
sample 
at ions 
nc er ta 
pauc it 
the pr 
d Ow d 
t sulf 

diffe 
ng add 
ion wi 
e oxid 

a imed 
ty for 
e of t 
pecies 
da tion 
ance e 

excha 
water . 



. 14 
te o 
nts 
, ra 
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nted 
that 
g se 
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) asked : 
xidation , 
on the 
tify the 
nts in 
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veral 
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tions of 

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minerals , 

organic 
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number 
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trapping 
topic 

t may be 
cess, 
r to 
oxygen 



154 



CONCLUSIONS 

Simplified stoichiometric equations 
describing pyrite oxidation, in 
combination with experimentally 
determined isotope parameters, define a 
field on a 6 18 0s versus <5 18 Ow diagram. 
Data from laboratory experiments, as 
well as data for samples from aerobic 
environments not affected by marine 
sulfate, were found to plot within 
this field. This suggests that the 
equations may present a reasonable, 
simplified description of the overall 
oxidation process. Calculations based 
on these equations indicate that the 
fraction of sulfate oxygen Os derived 
from water molecules ranges from about 
29 to 100% for the experiments, and from 
35 to 90% for the field samples. 

The available experimental results 
also suggest that some of the water 
oxygen may be incorporated in the 
sulfate as a result of exchange of 
oxygen isotopes between the water and 
one or more intermediate sulfur species. 
The possibility that oxygen derived from 
H-0 molecules can act as an oxidant in 
purely chemical conversions, tends to be 
discounted . 

The <5 Os values for samples from 
some of the field sites reflect the 
presence of sulfate of mar ine-evapor ite 
origin. In addition, the data for the 
Pine Point samples indicate the presence 
of sulfate derived from oxidation of HS 
which was generated earlier by bacterial 
sulfate reduction. Clearly, such 
background SOf must be characterized 
before <5 Os and <$ Ow data can be used 
to monitor sulfide oxidation in mining 
environments . 

ACKNOWLEDGEMENTS 

Field samples analysed in the 
current study were made available by 
Boojum Research Ltd., Toronto, Ontario. 
The authors thank Boojum Research Ltd. 
and the Natural Sciences and Engineering 
Research Council for supporting the 
study and the Stable Isotope Laboratory, 
respec t ively . 

LITERATURE CITED 

Chiba, H. and Sakai, H. 1985. Oxygen 
isotope exchange rate between 
dissolved sulfate and water at 
hydrothermal temperatures. 
Geochimica et Cosmochimica Acta 
49:993-1000. 



Holser, W.T., Kaplan, I.R., Sakai, 
and Zak, I. 1979. Isotope 
geochemistry -of oxygen in the 
sedimentary sulfate cycle. 
Chemical Geology 25:1-17. 



H 



Horlbe, Y., Shigehara, K., and Takakuwa, 
Y. 1973. Isotope separation 
factor of carbon dioxide-water 
system and isotopic composition of 
atmospheric oxygen. Journal of 
Geophysical Research 78:2625-26 29. 

Krouse, H.R. 1980. Sulphur isotopes in 
our environment. In: Handbook of 
Environmental Isotope Geochemistry, 
Vol. 1, The Terrestrial Environment 
(Eds. P. Fritz and J. Ch. Fontes), 
Elsevier Scientific Publishing Co., 
Amsterdam, pp. 435-471. 

Krouse, H.R. 1987. Relationships 
between the sulphur and oxygen 
isotope composition of dissolved 
sulphate. In: Studies on Sulphur 
Isotope Variations in Nature. 
International Atomic Energy Agency, 
pp. 19-29. 

Lloyd, R.M. 1967. Oxygen-18 

composition of oceanic sulfate. 
Science L56 1:1228-1231. 

Lloyd, R.M. 1968. Oxygen isotope 
behaviour in the sulf ate-wa ter 
system. Journal of Geophysical 
Research 73:6099-6110. 

Mizutani, Y., and Rafter, T.A. 1969. 
Oxygen isotopic composition of 
sulphates-Part 4: Bacterial 
fractionation of oxygen 
isotopes in the reduction of 
sulphates and in the oxidation 
of sulphur. New Zealand 
Journal of Science 12:60-68. 

Ralph, B.J. 1979. Oxidative reactions 
in the sulfur cycle. In: 
Biogeochemical Cycling of Mineral- 
Forming Elements (Eds. P. A. 
Trudinger and P.J. Swaine), 
Elsevier Scientific Publishing Co., 
Amsterdam, pp. 369-400. 

Rimstidt, J.D., Chermak, J. A., and 

Newcomb, W.D. 1986. The oxidation 
of sulfide minerals. Extended 
Abstracts, 5th International 
Symposium on Water-Rock 
Interaction, Int. Assoc. Geochem. 
and Cosmochem., pp. 471-474. 

Sasaki, A., and Krouse, H.R. 1969. 

Sulfur isotopes and the Pine Point 
mineralization. Economic Geology 
64:718-730. 

Schwarcz, H.P. and Cortecci, G. 1974. 
Isotopic analyses of spring and 
stream water sulfate from the 
Italian Alps and Apennines. 
Chemical Geology 13:285-294. 

Shakur, M.A. 1979. Oxygen and sulfur 
isotope studies of natural 
sulfates. M.Sc. Thesis, Department 
of Physics, The University of 
Calgary, 115 pp . 



155 



Smejkal, V. 1979. Oxygen isotopic 

composition of sulphates from some 
mineral waters and mine waters in 
Western Bohemia. International 
Atomic Energy Agency, Vienna, 
Isotope Hydrology 1978, Vol. 
1 :83-98. 

Taylor, B.E., Wheeler, M.C., and 

Nordstrom, D.K. 1984a. Stable 
isotope geochemistry of acid mine 
drainage: Experimental oxidation 
of pyrite. Geochimica et 
Cosmochimica Acta 48:2669-2678. 

Taylor, B.E., Wheeler, M.C., and 

Nordstrom, D.K. 1984b. Isotope 
composition of sulfate in acid mine 
drainage as measure of bacterial 
oxidation. Nature 308:538-541. 

Thode, H.G., Monster, J., and Dunford, 
H.B. 1961. Sulphur isotope 
geochemistry. Geochimica et 
Cosmochimica Acta 25:159-174. 

Ueda, A., and Krouse, H.R. 1986. 

Direct conversion of sulphide and 
sulphate minerals to SO2 for 
isotope analyses. Geochemical 
Journal 20:209-212. 

van Everdingen, R.O. and Krouse, H.R. 
1985. Isotope composition of 
sulfates generated by bacterial and 
abiological oxidation. Nature 
315:395-396. 

van Everdingen, R.O., Shakur, M.A.. and 
Krouse, H.R. 1982. 34 S and l8 
abundances differentiate Upper 
Cambrian and Lower Devonian 
gypsum-bearing units, District of 
Mackenzie, N.W.T. - An update. 
Canadian Journal of Earth Sciences 
19: 1246-1254. 

van Everdingen, R.O., Shakur, M.A., and 
Michel, F.A. 1985. Oxygen- and 
sulfur-isotope geochemistry of 
acidic groundwater discharge in 
British Columbia, Yukon, and 
District of Mackenzie, Canada. 
Canadian Journal of Earth Sciences 
22: 1689-1695. 

Weyer, K.U., Krouse, H.R., and Horwood, 
W.C. 1979. Investigation of 
regional geohydrology south of 
Great Slave Lake, Canada, utilizing 
natural sulphur and hydrogen 
isotope variations. International 
Atomic Energy Agency, Vienna, 
Isotope Hydrology 1978, Vol. 
1 :251-264. 



156 



GEOCHEMISTRY OF ABANDONED LIGNITE 
MINE SPOIL IN TEXAS 

Jan K. Horbaczewski and Frank Van Ryn 

Morrison-Knudsen Company, Inc. 



Abstract--Geochemical investigations of lignite 
mine spoils ranging in age from 10 years to over 35 
years were performed at an abandoned mined land 
reclamation project near Rockdale, Milam County, TX. 
A primary objective was to determine the degree to 
which oxidation of pyritic material had progressed 
for the design of an appropriate reclamation plan. 
For this purpose, eight continuous cores were 
recovered from the project area. The results 
indicated that within 10 years, consistent trends 
had already developed in the spoil among the princi- 
pal sulfur forms (organic sulfur, pyritic sulfur, 
and sulfate sulfur) . Three principal trends were 
identified: a gradual decrease in the proportion of 
sulfate sulfur to pyritic sulfur from the spoil 
surface downwards ; a change, marked by a transition 
zone which varied in location and definition accord- 
ing to the age of the spoil , from low pH values 
(less than pH 4.0) in near-surface spoil to higher 
pH values (greater than pH 5.0) at greater depths; a 
trend with age of spoil to the correlation between 
the pH transition zone and the proportion of sulfate 
sulfur. These trends, associated with the processes 
of oxidation and leaching, are depicted on triangu- 
lar graphs which are also used to show the recon- 
structed evolution of the spoil sulfur forms at 
ten-year intervals. The results of the geochemical 
investigations have been incorporated in the recla- 
mation plan which is designed to minimize the 
exposure, through regrading, of spoil still contain- 
ing pyritic sulfur forms. 

Additional Key Words: pyritic sulfur, acid- 
forming material, sulfur forms, vadose zone. 



157 



INTRODUCTION 

The work described in this paper was 
part of a reclamation program prepared for 
the Railroad Commission of Texas for the 
mitigation of an abandoned mined land 
reclamation project, near Rockdale, TX. 
Significant mining of lignite with drag- 
lines began in this area in the early 
1950' s and terminated in the project area 
with the passing of the Surface Mining and 
Reclamation Act of 1977. The spoils in the 
area , therefore , range in age from over 35 
years old to 10 years old. 

One of the principal objectives of the 
project was to determine the degree to 
which oxidation of pyritic sulfur had 
progressed and thus to assess the extent of 
acid and acid-forming materials for the 
reclamation plan (Railroad Commission of 
Texas, 1987). This paper deals specific- 
ally with the relationships discovered 
among the sulfur forms and pH values in the 
spoil . 



MATERIALS AND METHODS 

The data presented in this paper are 
based on eight continuous cores collected 
in the project area spoil. The project 
area covers approximately 1,200 acres and 
is located near Rockdale in Milam County, 
TX. The spoils are derived from the 
lignite-bearing formations of the Eocene 
Wilcox Group. 

Core sites were selected on the basis 
of geographic and geologic settings to be 
representative of the project area. Since 
there was no detailed information available 
on the premining overburden geology and 
geochemistry, the principal criteria in 
site selection were to represent the ' 4 
project subareas with an appropriate 
spatial distribution, to select sites 
showing a regular dragline spoiling pat- 
tern, and to distribute the sites in both 
geologically up-dip and geologically 
down-dip locations. 

At each site, a small bulldozer was 
first used to prepare a drilling platform 
on the spoil ridge. This resulted in a 
small loss of sample corresponding, on 
average, to the top 5 ft. of the spoil. 



Paper presented at the 1988 Mine 
Drainage and. Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and Reclamation 
and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement, April 
17-22, 1988, Pittsburgh, P. A.) 

2 
Jan K. Horbaczewski is Senior Envi- 
ronmental Specialist and Frank Van Ryn is 
Staff Environmental Specialist, Morrison- 
Knudsen Company, Inc., San Antonio, TX. 



All sample depths have been adjusted 
accordingly and are reported in relation 
to the original spoil surface. A truck- 
mounted, hollow-stem auger (3-inch inter- 
nal diameter) with a wire-line attachment 
for a Shelby tube was then used to collect 
the spoil samples. The hollow-stem system 
eliminated the potential contamination 
from drilling fluids or material falling 
in from the sides of the borehole. Cores 
were recovered to a depth approximately 4 
ft. below the expected final topography 
after regrading. The core lengths varied 
in depth from 19 . to 49 . 7 f t . A total of 
264.4 ft. of spoil core was collected. 

Each core was inspected and described 
in the field by an experienced geologist, 
sealed in plastic bags, labelled and 
boxed. Selection of sampling intervals 
was performed by an experienced soil 
scientist on the basis of the lithologic 
description and further examination of the 
core. A total of 51 sampling intervals 
were selected with a range of sample 
interval length from 0.6 ft. to 13.8 ft. 
and an average of 5.2 ft. 

The laboratory methods used for the 
analysis of the spoil samples were those 
recommended by the Railroad Commission of 
Texas (1985). Sulfur forms (organic 
sulfur, pyritic sulfur, and sulfate 
sulfur) were determined according to EPA 
method 3.2.6 (EPA 1978), which separates 
the sulfur forms by extractant: HC1- 
extractable sulfur (mostly sulfates), 
HNOo-extractable sulfur (mostly pyritic 
sulfur) , and non-extractable sulfur 
(mostly organic sulfur) . The extracted 
fractions were then submitted for sulfur 
analysis using an induction furnace and 
automatic sulfur titrator. The spoil 
reaction (pH) was determined on a 1:1 
soil: water extract according to Black's 
method 60-3.4 (Black 1965). 

A quality check was performed on the 
laboratory results testing for internal 
consistency with other geochemical para- 
meters (not described in this paper) . 

RESULTS AND DISCUSSION 

An examination of the proportions of 
sulfur forms (organic sulfur, pyritic 
sulfur, and sulfate sulfur) in the spoil 
showed that there were some general trends 
in the geochemistry with depth (Table 1). 
The principal trend is illustrated in 
spoil core GC-1 (fig. 1). Near the 
surface of the spoil [sample 1 (7.0-11.0 
ft)] the pyritic sulfur content is zero 
and the only sulfur forms present are 
organic sulfur and sulfate sulfur. At 
successively greater depths [sample 2 
(11.0-12.6 ft); sample 3 (12.6-16.4 ft); 
and sample 4 (16.4-24.0 ft)] the propor- 
tion of sulfate sulfur decreases while the 
proportion of pyritic sulfur increases. 
At still greater depths [sample 5 (24.0- 
31.2 ft)] the pyritic sulfur proportions 
are high and the sulfate sulfur propor- 
tions are low (generally below 15% of the 



158 



TABLE 1 - Contents of Sulfur Forms in Spoil Cores From Abandoned Lignite in Mine Spoil. 






Spoil 


Sample 




Total 


Sulfate 


Pyritic 


Organic 


Total 


Sulfate 


Pyritic 


Organic 




Core 


Number 


Depths 


Sulfur 


Sulfur 


Sulfur 


Sulfur 


Sulfur 


Sulfur 


Sulfur 


Sulfur 


pH 






(feet) 




I 








Z 






GC-1 


1 


7.0-11.0 


0.23 


0.14 


<0.01 


0.09 


100 


61 


... 


39 


3.8 




2 


11.0-12.6 


0.83 


0.37 


0.04 


0.42 


100 


45 


5 


50 


3.7 




3 


12.6-16.4 


0.32 


0.13 


0.07 


0.12 


100 


41 


22 


37 


3.6 




4 


16.4-24.0 


0.84 


0.10 


0.49 


0.25 


100 


12 


58 


30 


5.8 




5 


24.0-31.2 


0.10 


<0.01 


0.04 


0.06 


100 


— - 


40 


60 


6.2 




6 


31.2-34.3 


0.18 


0.01 


0.10 


0.07 


100 


6 


56 


38 


6.5 




7 


34.3-39.3 


0.24 


<0.01 


0.14 


0.10 


100 


--- 


58 


42 


5.9 




8 


39.3-45.0 


0.12 


0.01 


0.04 


0.07 


100 


8 


33 


59 


6.6 




9 


45.0-52.0 


0.33 


0.09 


0.14 


0.10 


100 


27 


43 


30 


4.6 




10 


52.0-56.9 


0.53 


0.02 


0.38 


0.13 


100 


4 


72 


24 


5.7 


GC-2 


1 


12.0-16.0 


0.26 


0.20 


<0.01 


0.06 


100 


77 


... 


23 


3.8 




2 


16.0-26.4 


0.09 


0.03 


0.01 


0.05 


100 


33 


11 


56 


4.4 




3 


26.4-31.9 


0.07 


0.03 


<0.01 


0.04 


100 


43 





57 


4.8 


GC-3 


1 


5.0-9.0 


0.19 


0.14 


0.01 


0.04 


100 


74 


5 


21 


3.7 




2 


9.0-10.6 


0.29 


0.23 


0.01 


0.05 


100 


79 


3 


17 


4.1 




3 


10.6-14.0 


0.51 


0.36 


0.06 


0.09 


100 


71 


12 


17 


3.3 




4 


14.0-18.1 


0.55 


0.28 


0.21 


0.06 


100 


51 


38 


11 


3.8 




5 


18.1-21.2 


0.39 


0.14 


0.17 


0.08 


100 


36 


44 


20 


5.6 




6 


21.2-29.0 


0.58 


0.05 


0.41 


0.12 


100 


9 


71 


20 


5.3 




7 


29.0-32.0 


0.77 


0.10 


0.53 


0.14 


100 


13 


69 


18 


5.5 




8 


32.0-35.0 


0.77 


0.01 


0.61 


0.14 


100 


2 


80 


18 


5.3 


GC-4 


1 


10.0-14.0 


0.07 


0.02 


<0.01 


0.05 


100 


29 


... 


71 


4.8 




2 


14.0-21.7 


0.04 


<0.01 


<0.01 


0.04 


100 


--- 


--- 


100 


4.9 




3 


21.7-23.0 


0.29 


0.21 


0.01 


0.07 


100 


72 


3 


25 


3.6 




4 


23.0-29.3 


0.04 


0.01 


0.02 


0.01 


100 


25 


50 


25 


4.4 




5 


29.3-31.3 


0.22 


0.12 


0.07 


0.03 


100 


55 


32 


13 


4.7 




6 


31.3-35.0 


0.12 


0.07 


0.01 


0.04 


100 


58 


8 


33 


4.5 




7 


35.0-36.0 


0.28 


0.12 


0.08 


0.08 


100 


42 


29 


29 


5.3 




8 


36.0-42.3 


0.12 


0.05 


0.03 


0.04 


100 


42 


25 


33 


4.7 




9 


42.3-48.3 


0.41 


0.07 


0.22 


0.12 


100 


17 


54 


29 


6.1 




10 


48.3-50.4 


0.09 


0.04 


0.01 


0.04 


100 


44 


12 


44 


5.2 


GC-5 


1 


8.0-12.0 


0.05 


<0.01 


<0.01 


0.05 


100 








100 


7.6 




2 


12.0-22.0 


0.05 


<0.01 


0.01 


0.04 


100 


--- 


20 


80 


7.5 




3 


22.0-32.2 


0.06 


<0.01 


0.02 


0.04 


100 


--- 


33 


67 


7.5 




4 


32.2-46.0 


0.06 


0.02 


<0.01 


0.04 


100 


33 


... 


67 


7.3 




5 


46.0-57.7 


0.05 


<0.01 


0.01 


0.04 


100 





20 


80 


7.4 


GC-6 


1 


4.0-8.0 


0.37 


0.28 


0.01 


0.08 


100 


76 


3 


21 


3.7 




2 


8.0-9.4 


0.78 


0.68 


<0.01 


0.10 


100 


87 


... 


13 


3.5 




3 


9.4-10.2 


0.27 


0.22 


<0.01 


0.05 


100 


81 


... 


19 


3.8 




4 


10.2-19.6 


0.35 


0.25 


0.03 


0.07 


100 


71 


9 


20 


4.1 




5 


19.6-27.3 


0.27 


0.11 


0.09 


0.07 


100 


41 


33 


26 


5.9 




6 


27.3-30.1 


0.14 


0.04 


0.04 


0.06 


100 


29 


29 


42 


6.5 




7 


30.1-35.9 


0.27 


<0.01 


0.19 


0.08 


100 





70 


30 


7.3 


GC-7 


1 


3.0-7.0 


0.04 


<0.01 


0.03 


0.01 


100 


--. 


75 


25 


7.0 




2 


7.0-16.0 


0.05 


<0.01 


0.02 


0.03 


100 


... 


40 


60 


5.2 




3 


16.0-24.0 


0.04 


0.03 


<0.01 


0.01 


100 


75 





25 


6.5 


GC-8 


1 


4.0-8.0 


0.51 


0.38 


0.04 


0.09 


100 


75 


8 


17 


3.5 




2 


8.0-12.8 


0.33 


0.26 


0.01 


0.06 


100 


79 


3 


18 


3.7 




3 


12.8-13.9 


0.06 


0.02 


<0.01 


0.04 


100 


33 


... 


67 


5.0 




4 


13.9-19.0 


0.23 


0.15 


0.01 


0.07 


100 


65 


4 


31 


7.0 




5 


19.0-24.0 


0.25 


0.09 


0.09 


0.07 


100 


36 


36 


28 


7.1 



159 



ORGANC S 







PYWT1C S 



SULFATE S 



GC-1 



GC-3 




GC-8 

3 - SAIvPLE NUMBER 



GC-1 

GC-3 

GC-8 

GC-6 

GC-4 



DATE OF SPOILING 

1 977 



1 975 
1 965 
1 960 
1 955 
1 950 



1 970 
1 965 
1 960 
1 955 



pyrtttc s 



GC— 6 SLLFSTE S 



GC-4 



FIGURE 1. SULFUR FORMS IN SPOIL CORES FROM ABANDONED 

LIGNITE MINE SPOIL. 



total sulfur forms). The trend is general, 
but there are exceptions , such as sample 9 
(45.0-52.0 ft.) which has an anomalously 
higher sulfate sulfur content. 

Associated with this trend is a 
secondary trend from low pH values (pH 3.6 
for sample 1) in the near-surface samples 
to higher pH values (pH 5.8 in sample 4) in 
the deeper samples (fig. 1). Spoil core 
GC-1 shows a trend in acidity from pH 3.8 
[sample 1 (7.0-11.0 ft)] decreasing to pH 
3.6 [sample 3 (12.6-16.4 ft)] and increas- 
ing to pK 5.8 [sample 4 (16.4-24.0 ft)]. 

A similar trend can be observed in 
spoil core GC-3 (fig.l) which has a pH 
value of 3.7 in near-surface spoil [sample 
1 (5.0-9.0 ft)] which decreases with depth 
to pH 3.3 [sample 3 (10.6-14.0 ft)] and 
finally increases to pH 5.6 [sample 5 
(18.1-21.2 ft)] at still greater depth. 
Similar trends can also be seen in cores 
GC-4, GC-6, and GC-8, although not so 
clearly. 

A third trend becomes apparent when 
the pH change is compared to the relative 
percentage of sulfate sulfur. In spoil core 
GC-1, pH values as low as 3.6 occur in a 
sample with sulfate sulfur proportions of 
41%; in spoil core GC-3 a pH value of 3.8 
occurs at a sulfate sulfur proportion of 
51%; and in spoil cores GC-4, GC-6, and 
GC-8, pH values less than 4.0 are associ- 
ated with proportions of sulfate sulfur 
greater than 70%. 



The first trend indicates that, even 
in spoil material which has been mixed by 
regular dragline side-casting, a vertical 
sequence in the proportion of sulfate 
sulfur to pyritic sulfur becomes estab- 
lished within approximately 10 years (e.g. 
spoil core GC-1). The results indicate 
that the highest proportions of sulfate 
sulfur to pyritic sulfur occur within the 
near-surface spoil and the ratio decreases 
with depth to the point where there is 
little or no sulfate sulfur. In the 
10-year-old spoil at GC-1, the absence of 
sulfate sulfur occurs at a depth of approx- 
imately 24 ft.; in spoil core GC-3 (20 
years old) , this point is reached at a 
depth of approximately 29 ft.; in spoil 
core GC-6 (30 years old) at over 30 ft.; 
and in spoil core GC-4 (35 years old) at 
over 48 ft. 

These depths are based solely on the 
proportions of sulfur forms and may be 
misleading because they do not take into 
account the effects of leaching. An 
appreciable sulfate sulfur content at a 
depth in the spoil may represent the 
in-place sulfate generated from pyritic 
sulfur oxidation, or it may represent the 
sulfate leached from the overlying spoil 
column above into a zone in which no 
significant pyrite oxidation has occurred, 
or it may represent a combination of both 
processes . 



160 



To determine the significance of 
leaching in the distribution of sulfate 
sulfur, the pH of the spoil nay be used as 
an indicator of pyritic sulfur oxidation. 
For example, in spoil core GC-1, the low pH 
values (less than pH 4.0) in sample 1 
(7.0-11.0 ft), sample 2 (11.0-12.6 ft), and 
sample 3 (12.6-16.4 ft) are indicative of 
active pyritic sulfur oxidation. The 
sample below sample 3 [sample 4 (16.4-24.0 
ft)] shows a value of pH 5.8, suggesting 
that oxidation has scarcely commenced. 
Similarly, based on pH values in spoil core 
GC-3, active oxidation of pyrite appears to 
be occurring to a depth of 18.1 feet 
(sample 4 - pH 3.8). The next lower sample 
(sample 5) has a pH value of 5.6. In spoil 
core GC-8, the transition to higher pH 
values occurs below 13.9 ft; in spoil core 
GC-6 below 19.6 ft; and in GC-4 below 23.0 
ft (although the distinction is not clear 
in this case) . 

If pH transitions are compared to the 
sulfate sulfur proportions, there appears 
to be a relationship that varies consis- 
tently with the age of the spoil. For 
example, the 10-year-old spoil in core GC-1 
has a pH transition zone associated with a 
sulfate sulfur proportion of approximately 
40%; the 20-year-old spoil in core GC-3 has 
a pH transition zone associated with a 
sulfate sulfur proportion of approximately 
501; and the spoil which is over 25 years 
old in cores GC-8, GC-6, and GC-4 has a pH 
transition zone associated with a sulfate 
sulfur proportion of over 70Z. 

If leaching of sulfate had occurred to 
any significant degree, then the greatest 
depletion of sulfate would be expected in 
the nesr-surface spoil samples where 
rainfall leaching would be most intense. 
Such a depletion of sulfate would be marked 
on the triangular graphs by an upswing of 
the near-surface spoil samples towards the 
organic sulfur apex, since these samples 
are also depleted in pyritic sulfur as 
shown earlier. Such trends are indeed 
exhibited by the older spoils. Spoil core 
GC-4, representing material that is over 30 
years old, shows the first two samples 
[sample 1 (10.0-14.0 ft.) and sample 2 
(14.0-21.7 ft.)] with no pyritic sulfur 
contents and very low sulfate sulfur 
contents. Similarly, spoil core GC-8 shows 
a near-surface sample [sample 3 (12.8-13.9 
ft.)] with no pyritic sulfur content and a 
moderately low (33%) sulfate sulfur con- 
tent. 

The spoil samples depleted of both 
pyritic sulfur and sulfate sulfur exhibit 
pH values that are higher than those 
associated with active oxidation of pyritic 
sulfur. In spoil core GC-4, sample 1 
(10.0-14.0 ft) shows a value of pH 4.8 and 
sample 2 a value of pH 4.9; in spoil core 
GC-8 sample 3 (10.6-14.0 ft) shows a value 
of 5.0. These pH levels indicate that 
either pyritic sulfur oxidation never 
occurred (because no pyritic sulfur was 
present in that depth interval) or that it 
has gone to completion and the original low 



pH has been raised by subsequent base 
resaturation. 

The distribution of sulfur forms in 
three other spoil cores provides further 
information on different geocheraical 
conditions to those already described (fig. 
2) . Spoil cores GC-2 and GC-7 have very 
low overall total sulfur contents and the 
sulfur forms were determined near the level 
of detection. Very small variations in 
content in any of the sulfur forms are 
therefore reflected in a very wide scatter 
on the triangular graphs. With the excep- 
tion of sample 1 in spoil core GC-2, these 
cores do not show significant reductions in 
pH due to pyritic sulfur oxidation. Spoil 
core GC-5 is atypical in that it displays a 
combination of very low total sulfur 
contents and high pH values (above pH 7.0) 
due to the probable presence of carbonates. 
The absence of pyritic sulfur and sulfate 
sulfur results in a cluster of data points 
on the triangular graph near the organic 
sulfur apex. 

SUMMARY AND CONCLUSIONS 

A number of general trends in the area 
associated with the age of the spoil can be 
observe that can be attributed to processes 
of sulfide oxidation and leaching. These 
may be examined using the proportions of 
the three sulfur forms (organic sulfur, 
pyritic sulfur and sulfate sulfur) and pH. 
From the available data, the hypothetical 
evolution of the spoil geochemistry in the 
vadose zone is reconstructed as follows: 

° In 10-year-old spoil (fig. 3) 
there would be a relatively clear 
sequence of samples with 
decreasing sulfate sulfur contents 
and increasing pyritic sulfur 
contents from the surface down. A 
transition zone would be observed 
in which there is a marked pH 
increase downwards from values 
less than 4.0 (indicating active 
and full oxidation of pyritic 
sulfur) to values greater than 5.0 
(indicating that oxidation of 
pyritic sulfur is, at best, 
commencing) . 

° After 20 years , the same samples 
would show a number of changes on 
the triangular graph. Since a 
large proportion of pyritic sulfur 
would have been oxidized, the 
entire string of data would be 
drawn closer to the sulfate sulfur 
apex. The pH transition zone would 
spread and would be less well- 
defined. Penetration of oxygen in 
the vadose zone would have allowed 
some oxidation of pyritic sulfur 
even in the deepest sample (No. 10) 
producing some sulfate sulfur. 
Some of the sulfate values in the 
lowermost samples might also repre- 
sent accumulations of material 
leached from above. 



161 



GC-2 



ORGANC S 




GC-5 



PYRTTTC S 



SULFATE S 




ORGANC S 



GC-7 




3 - SAMPLE NUvBER 



DATE OF SPOILING 

GC-2 1970 - 1975 

GC-5 1960 - 1965 

GC-7 1955 - 1960 



PYRTTTC S 



SULFATE S 



FIGURE 2. SULFUR FORMS IN SPOIL CORES FROM ABANDONED 

LIGNITE MINE SPOIL 



ORGANC S 



SURFACE 
SAM=LE 



SULFATE S 





20 



YEAR-OLD SPOIL 



JO - YEAR-OLD SPOIL 
FIGURE 3. RECONSTRUCTED EVOLUTION OF SULFUR FORMS IN SPOIL CORES 



162 



After 30 years, even more oxidation 
would have occurred and the top 5 
samples would show the pyritic sulfur 
proportion to be less than 10% of the 
total sulfur forms. The pH transi- 
tion zone would be very diffuse and 
difficult to define although it would 
be closer to the sulfate sulfur apex. 
Leaching of sulfate from the surface 
samples would have been significant. 



LITERATURE CITED 

Black, C. A. 1965. Methods of Soil Anal- 
ysis, Part 2. Chemical and Microbio- 
logical Properties, 1st Edition. 
American Society of Agronomy. (pp. 
920-923). 

Environmental Protection Agency. 1978. 

Field and Laboratory Methods Appli- 
cable to Overburdens and Minesoils. 
EPA-600/2-78-054. (pp. 60-62). 

Railroad Commission of Texas. 1987. 

Environmental Inventory Report for 
Abandoned Mined Land Project. Consul- 
tant's Report prepared by Morrison- 
Knudsen Company, Inc.. 

Railroad Commission of Texas. 1985. 

Topsoil Substitute Suitability Crite- 
ria; Material Suitable for Placement 
in the Top Four Feet of Levelled 
Minesoil; Overburden Parameters and 
Procedures - Update 3. Internal 
Memorandum, dated October 4, 1985. 



163 



THE RATE OF OXIDATION OF PYRITES FROM COAL AND ORE SOURCES 
AN AC IMPEDANCE STUDY 1 



S. Chander and A. Briceno 



Abstract. — The rate of oxidation of pyrites 
from coal and mineral sources has been determined 
using a new technique of AC impedance spectroscopy. 
In this technique the impedance of the pyrite/ 
solution interface is measured as a function of 
frequency from which the charge transfer resistance 
can be calculated. From the charge transfer 
resistance the rate of oxidation is calculated 
using the Stern-Geary equation. The results show 
that the rate of oxidation of the pyrite sample 
from a coal source is substantially greater than 
the corresponding rate of oxidation of the pyrite 
sample from an ore source. The charge transfer 
resistance of the ore pyrite increases with the 
extent of oxidation whereas the coal pyrite showed 
no such increase. In the case of ore pyrite the 
surface film acts as a passivating layer and 
retards the rate of oxidation. The formation of a 
protective film on coal pyrite is not observed. 



RATE OF PYRITE OXIDATION AND ACID MINE 
DRAINAGE 

The pollution of mine waters by 
acids generated through oxidation of 
pyrite is a well recognized problem, 
particularly for the coal industry. 
Although the potential for acid mine 
drainage (AMD) depends on a variety of 
factors which include sulfide content, 
sulfide form, sulfide surface area, rock 
type (acid neutralization capacity), and 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and Re- 
clamation and the U.S. Department of the 
Interior (Bureau of Mines and Office of 
Surface Mining Reclamation and 
Enforcement), April 17-22, 1988 , 
Pittsburgh, PA. 

S. Chander is Associate Professor of 
Mineral Processing and A. Briceno is a 
Graduate Assistant, Mineral Processing 
Section, The Pennsylvania State 
University, University Park, PA. 



rock permeability, prediction of the 
amount of acid generated still remains a 
difficult task (Caruccio 1968, Barton 
1978, Rose et al. 1983) - A major factor 
in the amount of free acid generated is 
the relative rates of acid generation 
through oxidation of the sulfide and the 
rate of acid consumption by the host 
rock. This is schematically illustrated 
in figure 1 in which G., and G„ refer to 



and C. and C 
rates of aci 



two different rates of acid generation 
refer to two different 
consumption. Clearly, the 
amount of acid generated, shown by the 
hatched area, is substantially greater 
for (G-, C.) conditions than it is for 
(Gp,CpJ conditions even though the total 
amount of sulfide is the same. Thus, one 
approach to decrease the AMD problem is 
to reduce the rate of oxidation of 
pyritic minerals. However, this approach 
requires an understanding of the factors 
that determine the rate of pyrite 
oxidation. 

The importance of the rate of 
oxidation of pyrite in AMD has been 
recognized by many investigators (Singer 



164 




10 



>!« C 2 ) by 
The hatchured areas 



TIME 

Figure 1. — A schematic representation of 
the amount of acid generated (G., or 
G ? ) by oxidation of sulfides ana the 
amount of acid consumed (C. 
the host rock, 
represent the amount of acid 
generated for two sets of 
conditions : 
[J]]- (G 1 , C 1 ) and g- (G 2 , C 2 ). 

and Stumm 1970, McKibben 1984), but the 
techniques for studying oxidation rates 
and the conditions of investigation vary 
considerably. The rates have been 
determined by: 

1) Measurements of the amount of acid 
generated (represented as acidity in 
CaCO., equivalent); 

2) Measurements of iron, species in 
solution as Fe and Fe ; 

3) Measurements of sulfur species in 

solution as SOT, S0~, etc; 
" x y 

4) Measurements of changes in pH or free 
H concentration in solution; 

5) Measurements of Eh of slurry which is 



a measure of Fe 
solution; 



/Fe 



etc 



6) Analysis of solid oxidation products 
(possible only when substantial fraction 
of pyrite has oxidized) . 

Although each of these techniques 
gives a measure of the rate of oxidation, 
the results cannot be directly related 
because pyrite oxidation occurs through 
formation of several intermediate ions 
like ferrous, ferric, thiosulfate, 
dithionate, etc.; and solid products like 
elemental sulfur, ferrous and ferric 
hydroxides, geothite, lepidocrocite , 
maghemite, and hematite. The study of 
the mechanism of pyrite oxidation, 
particularly in slightly acidic to 
slightly alkaline solutions, is further 
complicated by formation of surface coat- 
ings of the solid oxidation products. The 
presence of such oxidation products might 
change the rate limiting step from the 
chemical (or electrochemical) reaction 



AIR-DRIED 




MOSES ETAL. (1987) 

J 1_ 



20 



TIME ( minutes) 

Figure 2. — The amount of sulfate 

generated by pyrite oxidation in 
oxygen-saturated solutions at 
initial pH 9 (Moses et al. 1987). 

control to mass transfer control through 
the product layer. The results of pyrite 
oxidation reproduced in figure 2 from a 
recent study (Moses et al. 1987) clearly 
show that the data are difficult to 
intepret in terms of a kinetic model. 
Attempts by other investigators to 
intepret the simple kinetic data, parti- 
cularly for oxidation in mild conditions 
have also met with difficulties (Esposito 
et al. 1987). This investigation was 
undertaken to develop a new technique of 
AC impedance spectroscopy to measure 
reactivity of the pyrite/solution 
interface. 

AC IMPEDANCE SPECTROSCOPY 

The technique of AC impedance spec- 
troscopy to measure reactivity of pyrite 
is based on measurements of current 
through a pyrite electrode. The current 
is directly proportional to the rate of 
oxidation, which depends upon the poten- 
tial, E, of the electrode (the potential 
represents the driving force for oxida- 
tion). If a sinusoidal potential, 

E = AE sin <ut 

is applied, the current through the 
electrode is 

i = Ai sin (<ot + $) 

where AE is the amplitude of the 
AC signal, w is the angular frequency 
(=2irf, where f is the frequency in hertz), 
Ai is the amplitude of the current, and <(> 
is the phase lag. The impedance, Z, is 
defined as 

Z = dE/dl 



165 



cm 



-z* 

ohms 




Ro + Rp 



Figure 3. — a) A simple equivalent circuit 
for pyrite/solution interface. R - 
charge transfer resistance, C,, - 
interfacial capacitance, and R 
solution resistance, b) A Nyquist 
plot of equivalent circuit shown in 
figure 3a. 



In general, the impedance consists of a 
real component, Z', and an imaginary 
component, Z", such that 

Z = Z' + jZ" 

If the pyrite/solution interface is 
represented by a resistance term, R, and 
a capacitance term, C, 

Z' = R, 

Z" = 1/UC) 

Ac impedance spectroscopy involves 
measurements of the impedance as a 
function of frequency from which the 
values of R and C can be obtained. A 
simple equivalent circuit for the pyrite/ 
solution interface is given in Figure 3a 
where R represents the change transfer 
term foP the pyrite/solution interface, C 
represents the interfacial capacitance, 
and R represents the solution 
resistance. The results of the impedance 
measurements can be plotted in the form 
of a Nyquist plot shown in Figure 3b from 
which values of R can be obtained. The 
values of R obtained for pyrites from 
different sources are discussed in this 
paper. The capacitances were also 
obtained, but will be discussed else- 
where. The rate of oxidation of pyrite, 
which can be represented as the exchange 
current density, I , is related to the 
change transfer resistance through the 
Stern-Geary equation 



CRT 
DISPLAY 



WAVEFORM 
GENERATOR 



^^ 



POTENTIOSTAT 



DISC ORIVES 



LOCK-IN 
AMPLIFIER 



MICROPROCESSOR 




I/E CONVERTER 



PRINTER/ 
PLOTTER (Z. <jl) 



IMPEDANCE MEASUREMENT 
SYSTEM 



Figure 4. — A schematic of the impedance 
measurement system. W-working 
electrode (pyrite), C-counter 
electrode (graphite), and 
R-reference electrode. 



b b 
a c 



"o 2.303R 



+ b 



where b and b are the anodic and 
cathodic Tafel slopes, respectively. 



EXPERIMENTAL METHODS AND MATERIALS 

The impedance measurements were made 
with a Model 368-1 impedance system 
manufactured by EG&G Princeton Research. 
It consists of a Model 173 potentiostat 
and a Model 5706 lock-in amplifier. Both 
units are operated by an Apple lie 
microcomputer through a Model 276 
interface. Both the data acquisition and 
analysis were carried out with the aid of 
the computer. A typical impedance 
experiment consisted of the measurement 
of the magnitude and phase angle of the 
impedance as a function of frequency. An 
AC signal of lO^mV in the frequency range 
of 10 to 2x10 Hz was used in the 
impedance measurements. A schematic of 
the electrochemical setup is shown in 
figure 4. A three-electrode system was 
used in which the working electrode„was a 
polished specimen of pyrite of 1 cm 
geometric area. A saturated calomel 
electrode was used as the reference 
electrode and two graphite rods were used 
as the counter electrodes. 

The pyrite crystals were cut to 
size, and the exposed surface of the 
sample was polished on a 600-grit silicon 
carbide paper. The samples were polished 
wet under distilled water and then trans- 



166 



ferred to the cell quickly after washing. 
Samples were handled only by gloves dur- 
ing the entire polishing procedure. Al- 
though pyrite oxidation could not be pre- 
vented by this procedure, alternate chem- 
ical treatments were found to be unsatis- 
factory. The solution was deoxygenated by 
bubbling purified nitrogen overnight 
before inserting the sample in the cell. 
Nitrogen was bubbled throughout the con- 
ditioning period, which lasted about 
three hours unless otherwise stated. Dur- 
ing an electrochemical measurement, the 
flow of nitrogen at the surface of the 
solution was continued to eliminate dif- 
fusion of atmospheric oxygen into the 
cell. 

The ore pyrite sample was from Muro 
de Aguas, Logrono Province, Spain, and 
the coal pyrite was from a "sulfur ball" 
from the Lower Kittaning Seam, PA. 

Distilled water from a tin-lined 
Barnstead still (Model 210) equipped with 
a Q-baffle system was used in this 
investigation. Water of specific 
resistivity greater than 2 Mohm-cm was 
used in all the experiments, which were 
conducted at room temperature. 



RESULTS AND DISCUSSION 

The results of the AC impedance 
measurements for the coal and ore pyrite 
samples are presented in figure 5. The 
charge transfer resistance (R ) is plot- 
ted as a function of the DC pBtential for 
the pyrite/solution interface. It is 
small when the electrochemical reactivity 
of the interface is large (i.e., large 
reaction rate) and vice-versa. The 
results in figure 5 show that the coal 
pyrite is generally more reactive than 
the mineral pyrite in the potential range 
of -0.4 to 0.4 V. The functional depen- 
dance of the charge transfer resistance 
on potential is also different for the 
two pyrites. The ore pyrite sample 
shows a maximum in R at 0.25 V*. 
This potential is slSghtly greater than 
the rest potential of pyrite, which is 
0.19 V. In contrast to this behavior, 
the coal pyrite shows a maximum in the 
charge transfer resistance at 0.12 V, 
which is slightly below the rest 
potential of coal pyrite (0.28 V). The 
magnitude of the charge transfer 
resistance for this pyrite is about 25 
times smaller than the corresponding 
value for the ore pyrite. These results 
imply that the coal pyrite begins to 



POTENTIAL, 
-0.4 -0.2 



*The impedance measurements on ore 
samples from alternate sources show a 
similar behavior with the exception that 
the magnitude of the charge transfer 
resistance is different for each pyrite. 
The charge transfer resistance correlates 
very well with the oxidation behavior 
also. These results are published 
elsewhere (Chander and Briceno 1987). 




200 

150 
100 
50 



POTENTIAL, E„, V 

Figure 5. — The charge transfer resistance 
for an ore and a coal pyrite as a 
function of the DC potential. 



oxidize at lower potentials than ore 
pyrite. The exchange current densities 
obtained by using the Stern-Geary 
equation are plotted in figure 6 for the 
pyrites from the coal and the ore 
sources. The high reactivity of the coal 
pyrite is clearly seen. The filled 
symbols are the exchange current 
densities obtained by steady state 
polarization measurements. 

For the ore pyrite, the charge 
transfer resistance increases with 
increase in the extent of reaction as 
shown in figure 7, in which R is plotted 
as a function of charge transferred. In 
this series of tests, an anodic charge 
was passed to oxidize the pyrite under 
galvanostatic conditions (I.e. constant 
rate of oxidation), and the AC impedance 
measurements were carried out after a 
predetermined amount of charge was 
passed. The increase in R for the ore 
pyrite shows that a film forms on the 
surface of pyrite, the resistance of 
which increases with the amount of charge 
transferred or extent of the reaction. 
The film acts as a passivating layer and 
retards the rate of oxidation. A 
passivating film does not form on the 
surface of the coal pyrite, however. The 
results for coal pyrite, in figure 7, 
show that the charge transfer resistance 
does not increase with charge trans- 
ferred. The coal pyrite continues to 
oxidize rapidly due to the absence of a 
protective film. 

The nature of product layers on the 
surface of the reacted pyrites can be 
schematically represented as shown in 
figure 8. The product layer is 
relatively nonporous (or dense) at the 
surface of the ore pyrite (fig. 8a), 
whereas It is highly porous at the 
surface of the coal pyrite (fig. 8b). 



167 



Itf 



10' - 



10" - 



-0.6 



POTENTIAL, Ejct , V 
-0.4 -0.2 0.2 



0.4 





1 1 1 

COAL PYRITE 


1 


Oy 


1 


1 


- 


1 M SOO BORATE 
pH 9.3 


■a-g 






- 


1 


PYRITE 

1 1 


— A A 
1 


1 


1 





04 -0.2 0.2 0.4 
POTENTIAL, Eh, V 



0.6 



Figure 6. — The reactivities of pyrites 
from two different sources as a 
function of the DC potential. 



0.6 



0.5 - 



0.4 



0.3 - 



0.2 - 



GALV. CURRENT * 10 pA 
0.1 M SOD. BORATE 
pH 9.3 

ORE PYRITE 



COAL PYRITE 



50 IOC 150 200 250 

CHARGE, mC 



300 350 



400 



Figure 7. — The charge transfer resistance 
of ore and coal pyrites as a 
function of the extent of reaction 
(measured in terms of charge 
transferred) , 



Further analysis of the impedance data 
with the aid of a two-layer model 
(Chander et al. 1988) shows that the 
resistance of the coal pyrite/solution 
interface is primarily a charge transfer 
resistance (i.e., resistance of the 
porous film is negligible), whereas the 
resistance of the ore pyrite/solution 
interface consists of a charge transfer 
term for the product layer. Although the 
reasons for formation of a passivating 
film on the ore pyrite or the lack of a 
passivating film on the coal pyrite are 
not entirely clear, it is believed that 
the micropores in coal pyrite play an 
important role. For coal pyrite the 
dissolution occurs inside the pores or 
microcracks whereas precipitation of the 
product occurs outside the mouth of the 
micropore or crack. The product layer is 
sufficiently porous to allow diffusion of 
reactants and products. The conditions 



(a) Uniform Product Layer 
$57 



Unreacted 
Pyrite 



Solution 



^Re 



Reacted Layer 



( b) Non-uniform Product Layer 



Unreacted 
Pyrite 



Solution 



Reacted Layer 



Figure 8. — A schematic representation of 
surface layers on pyrite. 



which give rise to a change in the 
characteristics of the product layer can 
significantly alter the rate of pyrite 
oxidation. Additional studies are needed 
to determine the nature of surface films 
that form on the surface of pyrites and 
the role they play in pyrite oxidation 
phenomena . 



SUMMARY 

A new technique of AC impedance 
spectroscopy has been developed to 
determine the rate of oxidation of 
pyrite. This technique is especially 
useful for studying the reactivity of 
solids when product layers form at the 
interface. The study of oxidation of an 
ore pyrite, which is a hydrothermal 
pyrite, shows that a protective film 
forms on the surface of pyrite which 
retards the rate of reaction. In 
contrast, the microporous nature of coal 
pyrite, which is a brackish-water pyrite, 
gives rise to a nonproductive (highly 
porous) product layer. 



ACKNOWLEDGEMENTS 

The authors acknowledge the 
financial assistance from the National 
Science Foundation under grant, 
MSM-8413^77 and the U.S. Department of 
Energy under grant No. DE-FG22-85 PC80523 
in support of this research. 



LITERATURE CITED 

Barton, P. 1978, The acid mine drainage 
In Sulfur in the environment, Pt. 
II, Ecological Impacts. J.O. Nriagu 
(Ed.), Wiley, New York, NY, pp. 
313-358. 



168 



Caruccio, P.T., 1968, An evaluation of 

factors affecting acid mine drainage 
production and the ground water 
interactions in selected mine areas 
of Western Pennsylvania In Second 
Symposium on Coal Mine Drainage 
Research, Monroeville, PA, pp. 
107-152. 

Chander, S. and Briceno, A., 1987, Kine- 
tics of pyrite oxidation. Minerals 
and Metallurgical Processing, 
H: 171-176. 

Chander, S., Pang, J. and Briceno, A., 

1988, A two layer model for AC impe- 
dance analysis of pyrite/solution 
interface to be presented at the 
Symposium on Electrochemistry in 
Mineral and Metal Processing, the 
Electrochemical Society, Atlanta, 
GA, May 15-20. 

Esposito, M.C., Chander, S. and Apian, 
F.P., 1987, Characterization of 
pyrite from coal sources Iri Process 
Mineralogy VII. A.H. Vansiliou 
(Ed.). TMS/AIME, in press. 

McKibben, M.A., 1984, Kinetics of aqueous 
oxidation of pyrite by ferric iron, 
oxygen and hydrogen peroxide from pH 
1-H and 20-40°C. Ph.D. Thesis. The 
Pennsylvania State University, 159 
pp. 

Moses, CO., Nordstrom, D.K., Herman, 

J.S. and Mills, A.L., 1987, Aqueous 
pyrite oxidation by dissolved oxygen 
and ferric iron. Geochimica et 
Cosmochimica Acta. 51:1561-1571. 

Rose, A. W., Williams, E. G., and 

Parizek, R. B., 1983, Predicting 
potential for acid mine drainage 
from coal mines. Earth and Mineral 
Sciences Bull., The Pennsylvania 
State University 52:37-^1. 

Singer, P.C. and Stumm, W., 1970, Acid 
mine drainage: The rate limiting 
step. Science. 167:1121-23. 



169 



A QUANTITATIVE ASSESSMENT OF THE NATURAL ABATEMENT OF ACID MINE DRAINAGE BY 
INTERACTION WITH ALLEGHENY GROUP LITHOLOGIES 



Andrew D. Stahl and Richard R. Parizek 1 



Abstract. — Acid mine drainage is produced by the 
oxidation of iron sulfide minerals in the presence 
of air and water. Once this drainage enters a sat- 
urated ground-water system, oxidation is inhibited 
by the absence of air. The aquifer minerals will 
not be in equilibrium with the acid water and one 
or more reactions may proceed which results in acid 
drainage amelioration. In order to predict the 
rate and direction of migration of acid drainage 
contaminants within ground water, it is essential 
to be able to quantify the natural abatement capac- 
ity of the mine drainage affected rock. A lab- 
oratory procedure has been developed which may 
quantify the "acidity distribution coefficient" of 
acid mine drainage contaminated aquifers in which 
flow is dominated by fracture porosity. This 
technique is a modification of the batch adsorption 
tests commonly used to describe ion-exchange and 
adsorption reactions in porous media. All math- 
ematical models which describe the transport of a 
reactive contaminant must employ the distribution 
coefficient. For aquifers in which ground-water 
flow occurs primarily through fractures rather than 
intergranular pores, the distribution coefficient 
is a measure of the amount of contaminant adsorbed 
by the aquifer per surface area of rock exposed to 
the solution. Lower and Middle Kittanning cores, 
drill cuttings, and strip mine highwall samples 
have been analyzed for their capacity to ameliorate 
acidic solutions. All laboratory batch tests 
resulted in an increase in pH , regardless of the 
rock sample's acid-base account. All samples 
without carbonate material and with a low sulfur 
content were found to have a significant acid 
consuming capacity. All samples with an acid-base 
account greater than -40 (tons CaC03 equivalent per 
thousand tons of material) consumed acidity in the 
batch experiments. Some samples with an acid-base 
account greater than -140 but less than -40 
consumed acidity in the batch tests. Field tracer 



1 Andrew D. Stahl is a Research 
Assistant and Richard R. Parizek is a 
Professor of Geology, Department of 
Geosciences, The Pennsylvania State 
University, University Park, PA. 



170 



tests are another way of obtaining values for reac- 
tive contaminant transport parameters. Small scale 
acid-injection tests have been performed in wells 
from which the drill cuttings have been collected. 
The results of these tests illustrate that the 
developed laboratory technique has potential for 
predicting field occurrences. 



INTRODUCTION 

Acid mine drainage is produced by the 
oxidation of iron sulfide minerals in the 
presence of air and water. Once this 
drainage enters a saturated ground-water 
system, oxidation is inhibited by the 
absence of air. The aquifer minerals will 
not be in equilibrium with the acid water, 
and one or more reactions may proceed 
which results in acid drainage ameliora- 
tion. All mathematical models which 
describe the rate and direction of migra- 
tion of a contaminant that chemically 
interacts with the medium through which it 
flows must employ a parameter termed the 
"retardation factor". A reactive contami- 
nant, such as acid mine drainage, will 
migrate at a rate slower than that of the 
surrounding ground water. The retardation 
factor is a measure of the difference 
between these two flow velocities. Such 
delays in the transport of acid mine 
drainage have been noted in western 
Pennsylvania aquifers. Prior to the 
research presented in this paper, there 
had been no serious attempt to quantify 
acid drainage transport parameters. 
Hence, the ability to predict the 
migration of acid mine drainage within 
soils, aquifers, and confining beds had 
been very poor. 

Reactive Contaminant Transport Parameters 



Much work has 
quantifying the re 
contaminants other 
acid mine drainage 
often involve the 
distribution coeff 
discussed by Freez 
systems dominated 
the definition of 
coefficient and it 
retardation factor 
the equations: 



been conducted in 
tardation of reactive 

than those relating to 
These analyses most 
determination of the 
icient (Kd ) . As 
e and Cherry (1979) , in 
by intergranular flow, 
the distribution 
s relation to the 

(R) are described by 



Kd = (mass of solute on the solid phase 
per unit mass of solid phase) / 
concentration of solute in solution 

R = 1 + (bulk density/porosity) * Kd 

The method most often employed for Kd 
measurement is the laboratory batch 
experiment. These tests are performed by 
placing crushed rock in contact with a 
contaminated solution. The mixture is 
agitated on a shaker until equilibrium is 
reached. The effluent is then withdrawn 
and analyzed for the amount of contaminant 
retained by the rock. Knowing the initial 



and final solute concentrations, the 
volume of the solution, and the mass of 
the dry rock, the Kd can then be calcu- 
lated (Pickens et al. 1981). Another 
approach for Kd measurement is the field 
tracer test. This test is conducted by 
injecting the contaminated solution into a 
formation via a drillhole. After a 
certain amount of contact time with the 
rock, water samples are then collected and 
analyzed for the amount of contaminant 
retained by the formation. 

For systems dominated by fracture 
flow, Burkholder (1976) defines the 
distribution coefficient (Ka ) and the 
retardation factor: 

K a = (mass of solute on solid phase 

per unit area of solid phase) / 
concentration of solute in solution 

R = 1 + (ratio of surface area to 
void space * K a ) 

Previous Work 

The majority of previous work 
relating to chemical changes of acid mine 
drainage has been concerned with the 
effectiveness and practicality of various 
acid drainage treatment procedures. 
However, more closely related to the 
research presented in this paper, Crouse 
and Rose (1976) analyzed the natural 
ameliorative processes occurring in an 
acid mine drainage affected stream. They 
concluded that the stream sediments 
effected benef iciation by consuming 
hydrogen ions as a result of cation- 
exchange reactions and reactions with 
primary and secondary silicates. Ciolkosz 
et al. (1973) conducted soil percolation 
experiments and laboratory batch tests to 
analyze the ability of soils to renovate 
acid mine water. They discovered that 
maximum effect on pH, total acidity, 
conductivity, and iron content of the 
solution occurred within five minutes and 
in a 1:1 solid to solution ratio. They 
batch tested 21 Pennsylvania soils and 
learned that most of the variation of pH, 
total acidity, and iron were accounted for 
by cation-exchange capacity and CaC03 
equivalent . 

A time-distance acid-front migration 
finite element analysis is incorporated 
into the GEOFLOW model of D'Appolonia 
Waste Management Services, Inc. 
(Haji-Djafari 1983) . The determination of 
the acidity retardation factor in this 
model is based only on the neutralization 



171 



capacity of the porous media. Through a 
conversation with one of the GEOFLOW 
modelers it was learned that much diffi- 
culty exists in obtaining reliable results 
with this acid-transport routine (Snyder 
1986) . A discussion exists within the 
GEOFLOW user's manual suggesting the use 
of laboratory test data "to simulate acid- 
front migration and variations in solute 
retardation factors". 

Research Objectives 

One goal of this research is to 
determine whether a lithology completely 
lacking carbonate material, or otherwise 
possessing a negative acid-base account as 
determined by Environmental Protection 
Agency procedures (Sobek et al . 1978), can 
nevertheless consume acidity through ion- 
exchange and surface adsorption reactions. 

This question was addressed with the 
development of a laboratory procedure 
derived from the batch adsorption tests 
commonly used to describe ion-exchange and 
adsorption reactions in porous media 
(Relyea et al . 1980). Upon establishment 
of a reproducible batch test experimental 
method, the nature of contaminant trans- 
port in coal measure strata was consid- 
ered. It has been observed that much 
ground-water flow in the coal measures of 
western Pennsylvania occurs in rock with 
very little if any intergranular porosity. 
The main avenues for ground-water movement 
in these fracture-dominated flow systems 
are joints and bedding plane partings in 
which acid abatement reactions are limited 
by the available surface area of the 
fractures and the mineralogy exposed on 
these surfaces. For these reasons, both 
weathered and unweathered rock have been 
analyzed, and available surface area is a 
controlled parameter in the developed 
laboratory batch technique. Through 
additional research it is hoped that this 
laboratory procedure will be modified into 
a valid procedure for quantification of 
the natural acid abatement processes 
occurring in a fracture-dominated flow 
system. 

DESCRIPTION OF RESEARCH 

Rock Sample Collection and Preparation 

Figure 1 describes each rock sample 
used in the laboratory batch experiments. 
Weathered surfaces were broken off of the 
highwall samples and treated as separate 
samples. All samples except the drill 
cuttings were crushed by a jaw crusher. 
Representative splits were taken from each 
sample; part of the split was used for 
cation-exchange capacity tests. The rest 
of the split was pulverized for analyses 
including neutralization potential, maxi- 
mum potential acidity, x-ray diffraction, 
and carbonate carbon content. In prepara- 
tion for the batch experiments, the 
material not part of these splits was 
separated into size fractions by mechani- 
cal sieving with a nested sieve series and 



HIGHWALL SAMPLES : 

Random grab samples of Lower Kittanning 

overburden, Clearfield County, PA 



unwea the red 



weathered 




]ll- 6 ] 

unde relay 

AIR ROTARY DRILL-CUTTING SAMPLES : 
2-ft composite samples of Lower Kittaning 
overburden, dark gray to black shale, 
Clarion County, PA 



feet 
above 

L.K. 
Coal 



A-l 


B-l 


C-l 




E-l 


F-l 


A-2 


B-2 


C-2 


D-l 


E-2 


F-2 


A- 3 


B-3 


C-3 


D-2 


E-3 


F-3 


A-4 


B-A 


C-A 


D-3 


E-A 


F-A 



BCD 
DRILLHOLE 



Samples | H-Jl| and | H-Jdj are highly 
weathered shale samples. They were found 
filling joints of the Middle Kittaning 
highwall exposed above the strata 
penetrated by the Lower Kittanning 
drillholes. H-Jl is lighter in color than 
H-Jd. Both are rich in goethite. 



CORE SAMPLES : 

Each sample is 0.5 ft of core, Clarion 

County, PA 



Dark gray shale 

Black shale/boney coal 

Black shale with coal 

streaks 

Black shale with coal 

streaks 



Middle 

Kittanning 

overburden 



Lower 

Kittanning 

overburden 



Co- 


-1 


Co- 


-2 


Co- 


-3 


Co- 


-4 



Co-5 
through 
Co-14 



Dark gray shale 



Figure 1. — Description of rock samples 

used in laboratory batch experiments. 



172 



shaker. X-ray diffraction analyses were 
performed on different size fractions of 
highwall samples H-2; the material 
retained on sieve #4 was pulverized and 
its x-ray pattern was compared to that of 
the particles passing sieve #100 and the 
pattern of the initial representative 
split. All three patterns indicate 
identical mineralogy. 

Batch Test Procedure 

Each sample was analyzed for its 
natural acid abatement capacity through 
the following laboratory batch experiment 
procedure. Rock particles passing the #3 
mesh sieve (6.680 mm) and retained on the 
#4 mesh sieve (4.699 mm) are rinsed in 
distilled water to remove all silt and 
clay-sized material. The gravel is then 
thoroughly dried by compressed air. One 
hundred grams of the dry particles are 
placed in a 500 mL Erlenmeyer flask with 
100 mL of hydrochloric acid solution (pH = 
2.1). The mixture is gently agitated on a 
wrist-action shaker for 5 hours. The 
solution is then carefully poured from the 
rock and its pH, Eh, and temperature are 
measured. The sample is then filtered, 
and 50 mL are set aside for total acidity 
titration. The remaining solution is 
again measured for pH, Eh, and temperature 
as well as specific conductance. After 
these measurements are made the sample is 
acidified and stored for later analyses of 
cations and sulfate. At this time the 
total acidity titration is performed. The 
rock particles are again rinsed in dis- 
tilled water, dried, and stored. All 
water analysis procedures conform to the 
methods of the U.S. Environmental 
Protection Agency (1979) and/or the 
American Public Health Association (1980). 
The research which led to the development 
of this laboratory batch test procedure is 
discussed in the next section. 

At this point it should be stressed 
that at the present stage of research and 
development, the established laboratory 
technique can not predict field occur- 
rences. A calibration procedure would 
need to be developed in order to allow 
laboratory batch test results to be used 
in field transport models under variable 
conditions. The batch experiment proce- 
dure described above was derived to yield 
reproducible results that will address 
questions such as that stated earlier; 
i.e., whether lithologies with a negative 
acid-base account can consume acidity 
through ion-exchange and adsorption 
reactions. It is recognized by the 
authors that a hydrochloric acid solution 
at a pH of 2.1 does not realistically 
mimic an acid mine drainage plume. We do 
feel, however, that this is a logical 
starting point for continuing research. 

Derivation Of The Batch Test Procedure 

Highwall samples H-2 and H-3 were used 
for the trial and error experimentation 



which resulted in the procedure described 
above. These samples were chosen after 
preliminary rock analyses and batch tests 
produced favorable results. H-2, a gray 
shale, has no noticeable intergranular 
porosity. H-3, a slightly silty gray 
shale, has very little if any intergranu- 
lar porosity. They were found to have no 
carbonate carbon and very little sulfur. 
The batch tests resulted in significant 
increases in pH and decreases in total 
acidity and specific conductance. 

Many of the preliminary batch 
experiments were devoted to the careful 
control of available rock surface area. 
One major problem was prevention of par- 
ticle abrasion and rounding during the 
contact time. As expected, there was a 
direct relationship between the decrease 
in acidity and the amount of mud generated 
by abrasion. A very specific agitation 
level was found to be ideal for preventing 
abrasion; the shaking intensity on the 
wrist-action shaker was adjusted so that 
the solution remained slightly turbulent 
while the rock particles remained station- 
ary. Also contributing to the available 
surface area are silt and clay-sized 
particles adhering to the larger parti- 
cles. It was feared that briefly rinsing 
the particles in distilled water might 
cause desorption of ions thereby arti- 
ficially creating adsorption sites. 
However, batch tests conducted without 
rinsing resulted in a greater decrease in 
acidity than tests conducted with rinsing; 
the exposed surface area of the fine 
particles consumed more acidity than did 
artificially created adsorption sites. 
Therefore, in evaluating the amount of 
acidity consumed per unit surface area of 
rock, not rinsing the particles introduces 
more error than does rinsing. It was 
learned that drying the rinsed particles 
with heat caused them to become brittle 
and very easily abraded. Drying with 
compressed air caused no such problem. 

Once the above problems were resolved 
numerous batch tests were conducted on H-2 
and H-3 in order to determine a contact 
time, particle size, and initial solution 
concentration that would be reasonable and 
would yield a reproducible batch test 
procedure. The established values have 
been mentioned above. 

As discussed earlier, numerous rock 
and water analyses have been performed. 
However, considering the scope of this 
paper and space limitations, much of these 
data are excluded and will be published in 
the near future. This includes, in 
particular, cation-exchange capacities of 
rock samples, and batch-test solution 
sulfate and cation concentrations. The 
only data discussed in this paper are of 
those parameters found in table 1. 
Furthermore, the geochemical reactions and 
their kinetics deserve a much more 
elaborate discussion than can be presented 
here . 



173 




12 16 
Contact Time (hours) 




Figure 2. --Change in chemistry versus 
contact time, sample H-2. 



2 3 4 5 

(X 1000) 
Rook Surface Area < thous. of si. cm) 

Figure 4. — Change in chemistry versus rock 
surface area, sample H-2. 



Figures 2 and 3 illustrate the change 
in chemistry versus contact time for H-2 
and H-3 respectively. Particles retained 
on the #4 mesh sieve were equilibrated 
with a hydrochloric acid solution of pH = 
2.1 . Batch tests were conducted for 1, 
2, 3, 4, 5, 6, and 24 hours. Five hours 
was chosen as the contact time for the 
remaining tests. Figure 2 shows the 
decrease in hydrogen-ion activity contin- 
uing after 5 hours (the pH continued to 
rise) . In this case it appears 5 hours is 
a conservative choice. In contaminant 
transport considerations it is always 
better to underestimate the adsorption 
capabilities of an aquifer rather than 
overestimate it. A peak in the consump- 
tion of total acidity and conductance is 
indicated at about 6 hours. It may be 
that oxidation of the small pyrite content 
of H-2 may eventually become an influenc- 
ing factor. In natural ground-water 
systems there is much less available 
oxygen compared to the oxygen available in 
these batch adsorption experiments. 
Therefore, in aquifers, pyrite oxidation 




is usually limited or prevented. Data 
supporting this statement are presented 
later. As will also be discussed later, 
the occurrence of pyrite oxidation in the 
batch test versus in the aquifer is an 
aspect of this research that needs further 
investigation. Figure 3 shows a relative- 
ly constant pH and specific conductance 
after 5 hours. The continued consumption 
of acidity may be related to the slight 
intergranular porosity that H-3 may 
possess. Once again, it is more conser- 
vative and safe to assume only fracture 
flow when in actuality, some intergranular 
flow does occur. 

Figures 4 and 5 illustrate the change 
in chemistry versus rock surface area 
exposed for H-2 and H-3 respectively. The 
rock was equilibrated with a hydrochloric 
acid solution of pH = 2 . 1 . Batch tests 
were conducted for particles retained on 
mesh numbers 4, 5, 7, 10, 20, 30, and 45. 
The x-axis scale of these figures is 
thousands of cm 2 . This surface area was 
calculated assuming the particles to be 



1 1 L> 


— 1 — 1 — 1 — [—■ 

Hydrogen-Ion 


1 1 1 1 
Activity 

Total 


T — i — i — i — r- 1 — 1 1 '1 

Acidity "~~T~~~~~-~— « 














Specific 


Conductance 


V 









12 IS 

Contact Tine (hours) 

Figure 3. --Change in chemistry versus 
contact time, sample H-3. 



12 3 4 5 

(X 1000) 
Sock Surface Area ( thous. of sq. cm) 

Figure 5. — Change in chemistry versus rock 
surface area, sample H-3. 



174 




-200 

2.8 3.2 

Initial fH 

Figure 6. — Change in chemistry versus 
initial solution pH, sample H-2. 



spheres of diameter equal to the average 
of the mesh sizes of the sieve on which 
the particle was retained and the sieve 
the particle last passed. Also assumed 
was an average density of 2.7 g/cm 3 . In 
almost all cases consumption of acidity, 
hydrogen ions, and conductance increases 
with increasing surface area. The choice 
of the largest particles, those retained 
on #4 mesh sieve, is the most conserva- 
tive. It is not clear why H-2 did not 
continue to increase its consumption of 
hydrogen ions with increasing surface 
area. 

Figures 6 and 7 illustrate the change 
in chemistry versus initial concentration 
of the hydrochloric acid solution for H-2 
and H-3 respectively. Rock particles 
retained on the #7 mesh sieve were equil- 
ibrated with the solutions. Batch tests 
were conducted with initial pH values of 
2.1, 3.1, 3.9, and 5.4 . Once again, the 
y-axis scale is labeled "percent 
decrease". Negative values indicate 
levels of total acidity or hydrogen-ion 



1 1 1 1 1 1 1 1 1 1 1 1 1 1 

a —"" ®~ 


I 1 1 1 


Hydrogen- 1 on 


Activity 


-© 


* Total Acidity \ 































-10l 1 L 



_l_l I l_ 



J I I l_ 



2 3 4 5 

Initial pH 

Figure 7. — Change in chemistry versus 
initial solution pH, sample H-3. 



activity greater than the initial solution 
values. Both H-2 and H-3 consumed a large 
percentage of hydrogen ions from each of 
the test solutions. H-3 consumed larger 
percentages of acidity from the two 
solutions of higher initial concentration 
(low pH) than from the solutions of lower 
initial concentration. The same trend is 
seen for H-2 (fig. 6). The difference is, 
however, that H-2 produced acidity when 
equilibrated with the two solutions of 
lower concentration. For graphical 
clarity, the H-2 batch test data with the 
lowest initial concentration are not 
presented on figure 6. The test results 
are an increase of total acidity of 6.5 
times and a slight decrease in pH. These 
data suggest that batch tests conducted 
with solutions of low initial concentra- 
tion may not accurately describe field 
occurrences. As discussed above and 
below, acid production is suppressed in 
saturated ground-water systems. There- 
fore, the solution of highest initial 
concentration was chosen for the remaining 
batch experiments. 

The above discussion indicates the 
complications involved with using batch 
test analyses for predicting the total 
acidity of ground water in an acid 
drainage-contaminated aquifer. With 
further research this problem should be 
better understood. However, the data do 
not suggest the same problem with 
hydrogen-ion activity. Although this 
parameter will continue to be referred to 
as the "percent decrease in hydrogen-ion 
activity", it may also be valid to present 
the results as the "hydrogen-ion distribu- 
tion coefficient" as discussed in the 
introductory sections of this paper. 
Helfferich (1962) discusses the equation, 
known as the Freundlich isotherm, which 
describes ion-exchange equilibria. When 
considering surface area rather than rock 
mass, the equation can be written: 

log S = b log C - log K a 

where: S = (Initial Concentration - Final 
Concentration) / Surface Area 
C = Final Concentration 
Ka = Distribution Coefficient 
b = Slope of the log - log 
Adsorption Relation 

In order to calculate the distribution 
coefficient (K a ) from laboratory batch 
experiment results, it must be shown that 
the Freundlich isotherm is such that b=l. 
Figure 8 shows the linear regression of 
log S versus log C for hydrogen-ion activ- 
ity measurements of H-3 batch experiments. 
The slope of the regression line is 0.93 
with a correlation coefficient of 0.89. 
This suggests that the developed batch 
test procedure may be valid for quanti- 
tative measurement of the hydrogen-ion 
distribution coefficient. Once again, 
until more field and laboratory research 
is conducted, this parameter will continue 
to be referred to as the "decrease in 
hydrogen-ion activity". 



175 



9 -6.5- 



■ jr 

— i — i — i — i — I i i i i L_i ' i i I i i i i 



-5.6 
log C 



Figure 8. — Log-log adsorption relation 
with regression line of sample H-3 
batch test data. 



RESULTS AND CONCLUSIONS 

Batch Test And Field Tracer Test Results 

Table 1 contains rock analysis and 
batch test results for all samples shown 
on figure 1. All batch tests were 
performed with the technique described 
above . 

All highwall samples are found to be 
free of, or have very low levels of 
carbonate material. This includes the 
highly weathered joint-fill material 
collected from the Middle Kittanning 
highwall. H-l and H-lw are the only 
highwall samples which have a high poten- 
tial acidity as calculated from their 
sulfur content. With the exception of the 
sulfur-rich samples, the batch test 
results indicate each of the highwall 
samples to possess a significant capacity 
for consuming acidity. 



_ i i i i I i i i i | i r i » | i i i t |*« i p \ | i v «f i- 

• ■ 

a 

■ 

■ 

■ 
■ ' ' — 

■ : ■ 

■ : 

■ 
■ 

■ 
■ 

:■ 

■ 

■ 

■ 



20 



-220 -180 -140 -100 

Acid-Base Account 

Figure 9. — Decrease in H + activity versus 
acid-base account (all samples, 
account in tons of CaC03 equivalent 
per thousand tons of material). 



ii i i i i i i i 



| i i i i | i i i i | i rr i |T' i i | r n i | i i t i ii* i . r . 



nil I I I I I I I I I I I I I I I I I I I I I i i I i i i i I i i i I I I I 



■ i i i i i 



-160 -140 -120 -100 -80 -60 

Acid-Base Account 

Figure 10. — Percent decrease in total 

acidity versus acid-base account oi 
samples which consumed acidity 
(y-axis in hundreds of percentage 
points) . 



Figure 9 is a scatter plot of the 
percent decrease in hydrogen-ion activity 
versus rock sample acid-base account. All 
batch tests performed resulted in a 
decrease of hydrogen-ion activity, 
including batch tests where total acidity 
increased and the acid-base account was 
strongly negative. This adds more support 
for considering this a measurement of the 
hydrogen-ion distribution coefficient.- 



Figure 10 is a 
percent decrease in 
acid-base account, 
ed on this figure ar 
tests where total ac 
Acidity was consumed 
an acid-base account 
CaCOs equivalent per 
material. 



scatter plot of 
total acidity versus 
The only data present- 
e the results of batch 
idity decreased, 
by all samples having 
greater than -40 tons 
thousand tons of 



n 9.6 
t 



2.4 



:■ 

■ 
■ I 

. . J '■ m I : : 

■ 
: ■ ■ • 

■ 



-220 -190 -160 -130 -100 -TO 

ACid-Base Account 

Figure 11. — Percent increase in acidity 
versus acid-base account of samples 
which produced acidity (y-axis in 
hundreds of percentage points) . 



176 



Table 1. — Rock Sample characteristics and batch test results. 









Maximum 


flc id- 










Neu t r a lization 


Potential 


Base 


Percent 


Percent 


Percent 




Carbonate 
Carbon 


Potential 


Acidity 


Account 


Decrease of 
Hydrogen 


Decrease of 
Total 


Decrease of 


Sample 


(tons 


CaC03 equivalent per 


Specific 




</0 


thousand tons of 


material) 


[on Activity 


Acidity 


Conductance 


H-l 


0.000 


0.0 


150 


-150 


63 


-160 


-15 


H-2 


0.000 


.0 


5.7 


-5.7 


83 


60 


42 


H-3 


0.000 


1.5 


2.1 


-0.6 


92 


79 


63 


H-4 


0.000 


0.0 


1.1 


-1.1 


73 


57 


52 


H-5 


0.000 


0.0 


1.0 


-1.0 


53 


46 


38 


H-6 


0.001 


2.2 


1.1 


1.2 


93 


84 


66 


H-lu 


0.000 


0.0 


84 


-84 


60 


-150 


-5.1 


H-2u 


0.000 


0.0 


5.5 


-5.5 


87 


63 


30 


H-3u 


0.000 


2.2 


3.2 


-0.9 


97 


73 


68 


H-4u 


0.000 


0.0 


1.0 


-1.0 


74 


58 


56 


H-5u 


0.000 


.0 


1.0 


-1.0 


56 


47 


43 


H-Jl 


0.011 


4.0 


4.3 


-0.3 


99 


98 


69 


H-Jd 


0.017 


5.2 


6.2 


-1.0 


99 


97 


72 


A-l 


0.486 


8.7 


4.3 


4.4 


99 


99 


65 


fl-2 


0.000 


0.0 


140 


-140 


64 


-1200 


-110 


fl-3 


0.113 


0.0 


74 


-74 


84 


-730 


-89 


fl-4 


0.444 


10 


18 


-8.0 


99 


96 


60 


B-l 


0.277 


5.2 


4.0 


1.2 


100 


94 


61 


B-2 


0.264 


0.0 


43 


-43 


97 


-15 


0.5 


B-3 


0.002 


0.0 


200 


-200 


78 


-340 


-26 


B-4 


0.551 


6.0 


42 


-36 


99 


81 


47 


C-l 


0.293 


6.9 


3.3 


3.7 


99 


92 


62 


C-2 


0.653 


11 


10 


1.0 


99 


94 


51 


C-3 


0.249 


0.0 


160 


-160 


78 


-260 


-18 


C-4 


0.513 


5.7 


59 


-53 


99 


92 


57 


D-l 


0.460 


10 


3.4 


6.6 


100 


99 


63 


D-2 


0.514 


16 


25 


-9.0 


100 


100 


69 


D-3 


0.058 


0.0 


130 


-130 


66 


-940 


-98 


E-l 


0.59B 


18 


3.3 


15 


100 


98 


64 


E-2 


0.442 


6.9 


130 


-123 


98 


32 


35 


E-3 


0.469 


7.9 


150 


-142 


99 


84 


54 


E-4 


0.774 


20 


12 


8.0 


100 


99 


64 


F-l 


0.920 


24 


4.5 


20 


100 


99 


61 


F-2 


0.715 


21 


82 


-61 


88 


-99 


-29 


F-3 


0.517 


5.2 


74 


-69 


84 


-190 


-42 


F-4 


0.524 


5.4 


37 


-32 


99 


96 


48 


Co- 1 


0.724 


8.8 


5.2 


3.6 


99 


89 


70 


Co-2 


0.000 


0.0 


44 


-44 


54 


-340 


-22 


Co-3 


0.000 


0.0 


58 


-58 


62 


-150 


-0.1 


Co-4 


0.000 


0.0 


77 


-77 


68 


-48 


20 


Co-5 


1.274 


91 


75 


16 


100 


100 


46 


Co-6 


0.609 


42 


77 


-35 


100 


100 


39 


Co-7 


0.338 


31 


83 


-52 


100 


100 


28 


Co-B 


0.309 


31 


88 


-57 


100 


100 


33 


Co-9 


0.473 


42 


97 


-55 


100 


100 


41 


Co- 10 


0.287 


27 


98 


-71 


100 


100 


45 


Co- 11 


0.333 


32 


99 


-67 


100 


100 


30 


Co- 12 


0.127 


11 


120 


-109 


99 


37 


-1.6 


Co- 13 


0.001 


0.0 


150 


-150 


76 


-390 


-38 


Co- 14 


0.002 


0.0 


210 


-210 


79 


-170 


-6.3 



Figure 11 is a scatter plot of 
percent increase in total acidity versus 
acid-base account. The only data present- 
ed on this figure are the results of batch 
tests where total acidity increased. A 
comparison of this figure to figure 10 
shows that some batch tests of samples 
with acid-base accounts greater than -140 
and less than -40 resulted in a decrease 
in total acidity while others resulted in 
an increase in total acidity. Analysis of 
rock parameters not listed in table 1 did 
not clearly show which characteristics 
indicate whether a sample will increase or 
decrease acidity in the laboratory batch 
experiments. Further research is needed 
in this area. 



Figure 12 shows the acid-base 
accounting for all of the drill-cutting 
samples. Fifteen of the 23 samples are 
found to have a negative account. The 
average acid-base account of the 23 
samples is -53 tons CaC03 equivalent per 
thousand tons of material. Figure 13 
shows the percent change in total acidity 
for each drill-cutting sample. Positive 
numbers indicate percent decrease of the 
initial total acidity. Negative numbers 
indicate percent increase of the initial 
total acidity. Fifteen of the 23 samples 
consumed acidity. As mentioned earlier, 
each test resulted in a decreased 
hydrogen-ion activity. Field acid- 
injection tracer tests were conducted 



177 



-*o 
so 

-80 

-too 

-120 - 
-140 
-160 - 
-180 

-2O0 



cr 



m 



E2 f7ir/ 



n 






0Hw 



1 1 1 1 - 

H A2 A3 A4 



81 82 B3 84 C1 C2 C3 C4 01 D2 03 El E2 E3 E4 F1 

Rock Sample 



- 1 — i — r" 

F2 F3 F4 



Figure 12 . --Acid-base account of drill 

cutting samples (in tons of CaC03 per 
thousand tons of material). 



in the drillholes from which these samples 
were collected. The tests resulted in an 
approximate 50% decrease in both the 
hydrogen-ion activity and total acidity. 
These calculations consider the natural 
dispersion of the aquifer. Although the 
aquifer rock was found to have a negative 
acid-base account, and some drill-cutting 
batch tests resulted in significant acid 
production, these field tests suggest that 
the natural abatement processes dominate 
in an acid-contaminated aquifer. If 
pyrite oxidation does occur, it is of 
secondary importance in acid migration 
predictions. A detailed description of 
the field methods and the results is not 
presented here, but will be published in 
the future. 

Summary Statements 



may eventually provide the necessary 
information for modeling acid contaminant 
transport. 

(2) All laboratory batch tests 
resulted in an increase in pH, regardless 
of whether acidity was consumed or 
produced and regardless of the acid-base 
account. 

(3) All samples lacking carbonate 
material and having low sulfur content 
were found to have a significant acid- 
consuming capacity. 

(4) All samples having an acid-base 
account greater than -40 (tons CaC03 
equivalent per thousand tons of material) 
consumed acidity in the batch experiments. 

(5) Some samples having an acid-base 
account greater than -140 but less than 
-40 consumed acidity in the batch experi- 
ments. 

(6) Field tracer tests show that an 
aquifer can have a significant natural 
capacity to abate acid mine drainage even 
if it has an overall negative acid-base 
account and contains some rock which will 
produce acidity in laboratory batch 
experiments . 



REFERENCES 

American Public Health Association. 1980. 
Standard methods for the Examination 
of Water and Wastewater. 15th 
edition, 1134pp. A.P.H.A., New York, 
NY. 



(1) A laboratory technique has been 
established which may enable the quanti- 
fication of the hydrogen-ion distribution 
coefficient. With additional research, it 





10 




000 




-0 10 


1 

1 


-0 20 
-0 30 




c c 


-O 40 
-0 50 



\X YM__J\YM_ _YMY\_ YX^nYM—VX 



?o -060 

fife 

-0 70 

S -080 

• -090 

-1O0 

-1 10 

-120 



\AV 



2 



-i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — i — r 

»1 A2 A3 A4 81 82 B3 84 CI C2 C3 04 01 D2 D3 El E2 E3 E4 F1 F2 F3 F4 



Rock Sompta 

Figure 13. --Percent change in total 

acidity for drill-cutting samples 
(y-axis in thousand of percentage 
points. Positive no. indicates 
acidity consumed, negative no. 
indicates acidity produced) . 



Burkholder, H.C. 1976. Nuclide Migration 
for Faulted Media. In Nuclear Waste 
Management and Transportation 
Quarterly Progress Report October 
Through December 1975. Battelle 
Pacific Northwest Laboratories, 
BNWL-1978, Richland, WA. 

Ciolkosz, E.J., L.T. Kardos , and W.F. 

Beers. 1973. Soil as a Medium for 
the Renovation of Acid Mine Drainage. 
Research Project Technical Completion 
Report. 135pp. The Pennsylvania 
State University, Institute for 
Research on Land and Water Resources, 
University Park, PA. 

Crouse, H.L. and A.W. Rose. 1976. 

Natural Benefication of Acid Mine 
Drainage by Interaction of Stream 
Water with Stream Sediment. In Proc. 
6th Symposium on Coal Mine Drainage 
Research. pp. 237-269. National Coal 
Association. 

Freeze, R.A. and J. A. Cherry. 1979. 
Groundwater. 604pp. Prentice-Hall, 
Inc., Englewood Cliffs, NJ. 



178 



Haji-Djafari, S. 1983. GEOFLOW Ground 

Water Flow and Mass Transport Computer 
Program User's Manual. D'Appolonia 
Waste Management Services, Inc., 
Pittsburgh, PA. 

Helfferich, F. 1962. Ion Exchange. 

624pp. McGraw-Hill, Inc., New York, 
NY. 

Pickens, J.F., R.E. Jackson, and K.I. 
Inch. 1981. Measurement of 
Distribution Coefficients Using a 
Radial Injection Dual-Tracer Test. 
Water Resour. Res. 17 ( 3) : 529-544 . 

Relyea, J.F., R.J. Serne, and D. Rai . 
1980. Methods for Determining 
Radionuclide Retardation Factors: 
Status Report. Pacific Northwest 
Laboratory, U.S. Department of Energy, 
PNL-3349, Richland, WA. 

Snyder, M. 1986. Personal Communication. 
Pittsburgh, PA. 

Sobek, A. A. , W.A. Schuller, J.R. Freeman, 
and R.M. Smith. 1978. Field and 
Laboratory Methods Applicable to 
Overburdens and Minespoils. 
EPA-600/2-78-054. 203pp. U.S. 
Environmental Protection Agency, 
Cincinnati, OH. 

U.S. Environmental Protection Agency. 

1979. Methods for Chemical Analysis 
of Water and Wastes. EPA-600 
4-79-020. U.S. Environmental 
Protection Agency, Cincinnati, OH. 



179 



PREDICTION OF ACID MINE DRAINAGE FROM DULUTH COMPLEX 
MINING WASTES IN NORTHEASTERN MINNESOTA 

2 
Kim Lapakko 



Abstract. Ten Duluth Complex drill core 
samples, with sulfur contents ranging from 0.47 to 
2.17 percent, were leached in a laboratory experi- 
ment. The variation in sulfur content was due 
largely to fluctuations in iron sulfide content. 
Solids containing 0.8 percent sulfur or less pro- 
duced neutral drainage while, with one exception, 
solids with sulfur contents of 0.92 percent or more 
produced acidic drainage. The results indicated 
that acid production due to oxidation of sulfide 
minerals containing iron, such as pyrrhotite, 
increased linearly as the sulfur content increased. 
The acid consumption due to silicate mineral disso- 
lution was relatively constant despite minor 
variations in the silicate mineralogy. A sample 
containing 1.17 percent sulfur produced neutral 
drainage due to the buffering provided by a calcium 
carbonate content of three percent. Leaching of a 
sample from a test stockpile indicated that the fine 
fraction of stockpile solids exerted a major 
influence on the drainage chemistry. The pH range 
and temporal variation of drainage from solids 
containing 0.92 percent sulfur was similar to that 
from a test stockpile with a bulk sulfur content of 
0.6 percent. Analysis of the test stockpile solids 
indicated that particles smaller than 2.0 mm in 
diameter comprised about 12 percent of the stockpile 
mass and had a sulfur content of 1 percent. The 
laboratory results suggest these small particles 
exert a major influence on stockpile drainage 
quality, most likely due to their elevated sulfur 
content and specific surface area. 



INTRODUCTION 



The Duluth Compl 
northeastern Minnesot 
formation containing 
nickel sulfides (List 
1977). These deposit 
of the United States' 
the nation's largest 
resource (Minnesota E 
Board 1980, Kingston 



ex, located in 

a, is an intrusive 

low-grade copper and 

erud and Meineke 

s comprise 25 percent 

copper resource and 
nickel sulfide 
nvironmental Quality 
et al. 1970). 



Presented at the 1988 Mine Drainage 
and Surface Mine Reclamation conference 
sponsored by the American Society for 
Surface Mining and Reclamation and the 
U.S. Department of the Interior (Bureau 
of Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 
1988, 2 Pittsburgh, PA. 

Kim Lapakko is Senior Engineer, 
Minnesota Department of Natural Resources, 
St. Paul, MN. 



Development of these resources would 
require open pits which will eventually 
fill and discharge water, possibly contain- 
ing elevated concentrations of trace 
metals, to the environment (Thingvold et 
al. 1979). 

Based on conceptual models, the open „ 
pits could produce an estimated 4 to 10 km 
of waste rock stockpiles (Sturgess 1981; 
Veith 1978), which would generate an 
estimated runoff of 1.5 to 3.8 billion 
liters in a year of average precipitation 
(Hewett 1980). Trace metal concentrations 
in runoff from waste rock stockpiles at the 
Erie Mining Company Dunka Mine were 10 to 
10,000 times natural background levels of 
streams in the area (Eger et al. 1981). 
Since these metals have been shown to be 
toxic at low aqueous concentrations (United 
States Environmental Protection Agency 
1976), means of mitigating the impacts of 
mining drainage will be required for these 
resources to be developed. 



180 



The degree of mitigation required may 
be a function of the potential for trace 
metal and acid release from a given 
stockpile. This potential is dependent 
upon, among other factors, the composition 
of the stockpiled rock. Examining 
dissolution with respect to all chemical 
and mineralogical variables would, at 
best, be extremely complicated. Character- 
izing mine wastes with respect to these 
variables in the course of operations is 
also impractical. Consequently, a more 
simplistic course of experimentation, 
using sulfur content as the independent 
variable, was adopted. 

The effect of sulfur content on the 
leaching of Duluth Complex mining wastes 
has been examined for a limited number of 
samples. Field tests were conducted on 
test stockpiles constructed of rock 
removed from a test shaft. Over a 6-year 
period, test stockpiles containing 0.8 and 
1.4 percent sulfur produced acid leachate 
while piles containing 0.6 percent sulfur 
did not (Eger and Lapakko 1985). Subse- 
quent data indicate the drainage pH has at 
times decreased below pH 5 at one of the 
four piles containing 0.6 percent sulfur. 
In shake flask tests, neutral leachate was 
generated by tailings containing 0.92 
percent sulfur or less, while acidic 
leachate was generated by tailings with 
sulfur contents equaling or exceeding 2.65 
percent (Natarajan and Iwasaki 1983). 



The rate o 
has been report 
tional to the r 
surface area av 
Nelson 1978; Sa 
Mooney I960), 
surface area of 
function of the 
rock, particle 
the surface. Si 
content of mine 
sulfur content 
sulfide mineral 
importance due 
production. 



f metal sulfide oxidation 
ed to be directly propor- 
eactive sulfide mineral 
ailable (Lapakko 1980; 
to 1960a, 1960b; Sato and 
In turn, the available 
nonporous solids is a 
sulfide content of the 
size, and the roughness of 
nee the copper and nickel 

wastes is small, the 
is an indicator of iron 
s, which are of major 
to their role in acid 



Since other sulfides are present in 
limited quantities, their influence is 
diminished and to some extent uniform. By 
using uniformly small solids, the 
influence of particle size and host rock 
association of the iron sulfides is 
normalized. If the sulfides are liberated 
in small size fractions, the sulfur 
content will represent the available 
sulfide surface area, assuming specific 
surface area does not vary greatly among 
sulfide minerals of the same particle 
size. Since all the sulfide surface area 
is known to be available, acid production 
rates per unit sulfide surface area can be 
determined. Using this normalized value, 
the acid production rate can be extrap- 
olated to mining wastes based on their 
available sulfide surface area. This 
extrapolation will allow appropriate 
mitigation based on the rock composition. 



The use of sulfur content as the 
independent variable neglects the effects 
of many other variables, environmental as 
well as compositional, on the leaching of 
mine wastes. Although there is a given 
degree of compositional uniformity within 
an ore body, there are deviations from the 
norm. To accurately assess the leaching 
potential of mining wastes from a given 
body, a large number of samples must be 
analyzed. This implies the use of 
efficient and accurate laboratory techni- 
ques from which results can be extrapolated 
to operational conditions. This study 
involved examination of the leaching 
behavior of a small number of Duluth 
Complex samples with variable sulfur 
contents, and determination of the feasibi- 
lity of extrapolating laboratory results to 
operational conditions. 



METHODS 

Materials 

An experiment was conducted to examine 
the effect of solid phase sulfur content on 
the quality of drainage from Duluth Complex 
mining wastes. The sulfur content of the 11 
samples examined ranged from 0.47 to 2.17 
percent. Acid digestion (U.S. Bureau of 
Mines 1980) and subsequent analysis re- 
vealed typical solid phase copper and 
nickel contents of 0.13 to 0.23 and 0.05 to 
0.09 percent, respectively (table 1). The 
samples were selected to approximate mining 
wastes, based on a 0.2 percent copper 
cutoff. Mine waste would, of course, be 
larger with a maximum diameter of approxi- 
mately 45 cm for under-ground mines and 
120 cm for open pits. Plagioclase, 
olivine, pyroxenes, and biotite were the 
predominant minerals, as determined by 
examination of polished thin sections with 
an optical microscope (table 2). Specific 
surface area was determined by single point 
BET nitrogen adsorption by A. U. Baes of 
the Soil Science Department at the 
University of Minnesota, St. Paul. 

With one exception, the solids were 
MINNAMAX core samples which had been 
drilled from 1974 to 1977. The core 
samples had been crushed to minus 1/4 inch, 
placed into plastic bags, and stored inside 
55-gal barrels at the MINNAMAX site near 
Babbitt, MN. Under the direction of Karl 
Smith, the samples were processed for 
experimental use at the Mineral Resource 
Research Center at the University of 
Minnesota. The plus 10-mesh fraction was 
selected for experimentation in order to 
minimize the amount of surface area exposed 
to oxidation while the samples had been 
stored after drilling. This fraction was 
stage crushed using a pulverizer and wet- 
screened to segregate the -100/+270 mesh 
fraction for use in laboratory tests. 

The remaining sample (1.26 pet S) was 
collected from a test stockpile which was 
constructed in April 1977 at the MINNAMAX 



181 



Table 1. — Elemental composition of 11 Duluth Complex samples, values in weight percent 

























Pet of 












S with 












total S 












FeS 1 












bound by 


s 


Cu 


Ni 


Co 


Zn 


Fe 


Ca 


Mg 


Na 


K 


Fe 


0.47 


0.23 


0.060 


0.011 


0.017 


0.31 


15.2 


7.8 


7.9 


2.6 


0.37 


65 


.59 


.13 


.072 


.013 


.0070 


.44 


11.5 


11 


7.1 


3.1 


.27 


75 


.80 


.19 


.067 


.011 


.010 


.66 


13.7 


9.8 


6.6 


3.0 


.42 


82 


.92 


.091 


.049 


.014 


.016 


.83 


24.6 


8.0 


11.8 


2.2 


.42 


90 


1.17 


.23 


.082 


.011 


.015 


1.0 


16.3 


14.6 


10.5 


1.4 


.42 


85 


1.24 


.19 


.066 


.012 


.013 


1.1 


15.4 


6.8 


10.1 


2.3 


.42 


91 


1.35 


.23 


.086 


.018 


.011 


1.2 


18.4 


9.4 


7.3 


2.4 


.30 


87 


1.87 


.17 


.088 


.021 


.011 


1.7 


19 


9.6 


8.7 


7.5 


.86 


92 


2.01 


.20 


.055 


.012 


.020 


1.9 


19.6 


5.3 


5.1 


2.2 


1.10 


93 


2.17 
1.26 Z 


.20 


.078 


.010 


.027 


2.0 


8.5 


1.7 


2.9 


2.6 


1.83 


93 


.55 


.20 


.018 


.014 


.97 


18 


8.0 


8.0 


2.7 


.44 


77 



Calculated as the sulfur not bound by Cu, Ni, Co, or Zn assuming all trace metals were 
, bound by sulfur in a 1:1 mole ratio of sulfur to metal. 
"Sample from test pile. 



Table 2. — Approximate modal compositions of Duluth Complex Samples. 

(By Tatiana Sabelin, Mineral Resources Research Center, University of Minnesota) 

Mineralogy (%) 



PCT 
S 



Olivine Pyroxene Plagioclase 



Biotite 



Opaque Amphibole Other Calcite 



0.47 


30 


10 


45 


.59 


17 


4 


68 


.80 


20 


8 


59 


.92 


31 


27 


24 


1.17 


19 


15 


28 


1.24 


17 


27 


47 


1.35 


10 


14 


57 


1.87 


18 


4 


57 


2.01 


7 


15 


59 


2.17„ 
1.26 Z 


15 


8 


51 


15 


11 


49 



5 


8 


2 


5 


4 


7 


2 


16 


2 


8 


1 


5 


- 


9 


2 


10 


6 


13 


6 


17 


3 


16 



2 
4 
2 

25 

3 

10 

9 

3 

4 



^Includes plagioclase and quartz; of this total maybe as much as 1/3 to 1/2 is quartz, 
'Sample from test pile. 



site. The pile was 
mined during the de 
shaft. When part o 
removed in Septembe 
sample, weighing ov 
collected for analy 
distribution, speci 
chemistry, and mine 
1986). A subsample 
in the -100/+270 me 
collected for use i 



constructed of rock 
velopment of a test 
f the test pile was 
r 1982 a representative 
er 3 tons, was 
sis of particle size 
fie surface area, 
ralogy (Lapakko et al. 
of stockpile material 
sh fraction was 
n the experiments. 



Leaching Procedure 

Several techniques have been used in 
laboratory leaching of mining wastes 
(Caruccio 1986; Ferguson and Mehling 
1986). The method used was based on the 



principle that sulfide minerals oxidize in 
the presence of atmospheric oxygen and 
water (Gottschalk and Buehler 1912; 
Caruccio et al. 1980). Samples (75 g) of 
-100/+270 mesh rock were placed into the 
upper segment, or reactor, of a two-stage 
filter unit and rinsed with 200-mL volumes 
of distilled deionized water. Seven of the 
solids were run in duplicate while single 
reactors were used for the remaining four 
solids (0.47, 0.59, 1.35, and 2.01 pet S). 
The rinse water was allowed to remain in 
contact with the solids for five minutes 
and filtered through a 0.45 micrometer 
filter. Between rinses the solids were 
kept in the upper segment and stored in a 
box to dry. A cover was placed about 3 cm 
above the upper edge of the box to allow 
drying of the solids and prohibit the input 
of airborn debris. A thermostatically 



182 



controlled heating 
the box to maintain 
ture. Water contai 
box in an attempt t 
constant humidity, 
relative humidity i 
tored four to five 
average temperature 
standard deviation 
corresponding value 
were 56 and 9.1 per 



pad was placed beneath 

a constant tempera- 
ners were placed in the 
o maintain a fairly 

Temperature and 
n the box were moni- 
times per week. The 

was 25 C with a 
of 1.0 C, while the 
s for relative humidity 
cent . 



At the beginning of the experiment, 
the solids were rinsed five times to 
remove oxidation products generated after 
crushing and wet sieving. After this 
initial washing the solids were rinsed 
weekly. Four rinses were used after one 
week and three rinses after two weeks. 
Two rinses were found to remove the 
majority of sulfate from the solids and 
were used in subsequent weeks. 

The rinse water, or drainage, was 
analyzed for specific conductance, pH, 
sulfate, copper, nickel, cobalt, zinc, 
calcium, magnesium, sodium, and potassium. 
Specific conductance was analyzed using a 
Myron L conductivity meter, while a 
Radiometer 29 pH meter was used for pH 
analyses. Sulfate was analyzed using 
either the barium sulfate turbidimetric 
technique (APHA et al. 1975) or a 
Technicon autoanalyzer . Metals were 
analyzed with a Perkin Elmer 603 atomic 
absorption spectrophotometer. The experi- 
ments were conducted at the site 
previously leased by MINNAMAX, which is 
presently leased by Kennecott Minerals 
(fig. 1). 



RESULTS AND DISCUSSION 

Laboratory Data 

Drainage pH generally decreased as 
the solid phase sulfur content increased 
and as the duration of leaching increased. 
The pH of drainage from solids containing 
less than 0.9 percent S was consistently 
above 6, while drainage from solids 
containing more than 1.3 percent S was 
typically below pH 6. Between 0.9 and 1.3 
percent S was a transition zone. Drainage 
pH from one of the three transitional 
solids continuously remained above 7.5 
(fig. 2). The drainage pH from the other 
two transitional solids was initially in 
the neutral range, but steadily decreased 
to about 4.2 where it plateaued. The 
tendency for drainage pH to decrease and 
level over time was observed with most 
solids, but the decrease was most pro- 
nounced with these transitional solids. 
These pH trends were the net result of 
acid production by the oxidation of iron 
sulfide minerals and acid consumption by 
the dissolution of silicate minerals. 

The dissolution of iron sulfide 
minerals leads to the production of acid, 
as indicated by reaction 1 (Nelson 1978). 




Figure 1. — Project site in northeastern 
Minnesota. 



8 - 



7 - 



6 - 




O »" 1- T- 1- 



PERCENT SULFUR IN SOLID 



Figure 2. — Drainage pH versus sulfur 

content of the solid phase. Horizonal 
lines represent 25, 50 and 75 per 
percentiles, as well as the minimum 
and maximum values during the entire 
experiment . 



The iron sulfide oxidation may be 

FeS(s) + (3/2)H 2 + (9/4)0 2 = 

FeOOH(s) + 2H + (aq) + S0 4 2 "(aq) (1) 

influenced by microbial activity (Natarajan 
et al. 1982) and galvanic interactions with 
other sulfide minerals (Natarajan and 
Iwasaki 1983), as well as other factors. 
The oxidation of sulfide to the hexavalent 
sulfur in sulfate does not contribute acid. 
The two moles of acid produced are the net 
result of the oxidation of ferrous iron 



183 



and the subsequent precipitation of ferric 
iron, as lepidocrocite for example (Nelson 
1978; Sung and Morgan 1980). 

The rate of iron sulfide oxidation 
was reflected by the appearance of sulfate 
in solution. In an oxidizing environment, 
sulfate is the ultimate reaction product 
of sulfide oxidation. Over the aqueous 
concentration range observed in the 
laboratory samples, sulfate will remain in 
solution as opposed to precipitating. 

The hypotheses that sulfate was the 
dominant oxidation product and was stable 
in solution were supported by the fact 
that sulfate concentrations in stored 
samples were constant over several months. 
If reduced sulfur species were present 
they would have oxidized over this period 
and increased the sulfate concentration. 
Similarly there was no decrease in sulfate 
concentration which would result if 
sulfate precipitated from solution. The 
fact that more than 90 percent of sulfate 
concentrations were less than 50 mg/L, 
with none exceeding 130 mg/L, further 
indicates that sulfate precipitation or 
adsorption was unlikely. Some sulfate, or 
other sulfide oxidation products, may have 
been adsorbed onto the solids in the 
reactor. However, the multiple rinses 
indicated that little sulfate was removed 
from the solids by distilled water after 
the two rinses. 

Since sulfide was oxidized to sulfate 
and aqueous sulfate was stable, the rate 
of sulfide oxidation equals the rate of 
appearance of sulfate in solution at 
steady state. Since 82 to 93 percent of 
the sulfide present in these samples was 
bound by iron (table 1), oxidation of 
these iron sulfides was most likely the 
major sulfate source. Pyrrhotite is the 
predominant iron sulfide in the Duluth 
Complex (Bonnichsen 1972), and in these 
samples . 

The rate of sulfate appearance was 
proportional to the mean sulfate concen- 
tration, since the rinse volumes were 
essentially equal, the mass of solids 
leached was constant at 75 g, and the 
rinse interval was constant. Mean concen- 
trations from weeks 6 through 17 were used 
since a) sulfate concentrations during the 
initial 5 weeks were typically low and not 
representative of the sulfate release in 
general, and b) after the initial 17 
weeks, the rinsing interval for some of 
the solids was extended beyond the stand- 
ard duration of 1 week, which affected the 
rate of sulfate release. The weekly 
sulfate concentrations were fairly con- 
stant over this period, suggesting the 
oxidation of sulfide to sulfate had 
approached steady state. However, some of 
the solids subsequently generated sulfate 
concentrations which fluctuated from this 
narrow range. Although a steady-state 
condition may not have been reached for 
all solids, the sulfate data provide the 
best approximation of the sulfide 



oxidation rate during the initial stage of 
experimentation. The phrase "initial 
apparent steady-state period" best 
describes this phase. 

The rate of sulfide oxidation, as 
reflected by mean sulfate concentrations, 
correlated highly with the solid phase 
sulfur content. The percent sulfur in the 
solid phase accounted for 77.6 percent of 
the variation in the sulfide oxidation rate 
(r = 0.776). The r value increased to 
0.963 when two outlying data points at 1.35 
and 1.87 percent S were ignored. 

The degree of correlation for the 
entire data set increased when variations 
in specific surface area among the solids 
were taken into account. The anomalous 
sulfate concentrations, from the solids 
containing 1.35 and 1.87 percent S, were 
higher than predicted by the linear model 
describing the eight remaining points 
(fig. 3). Although all solids were in the 
-100/+270 mesh size fraction, limiting 
particle diameter to the range of 0.053 to 
0.149 mm, there were minor variations in 
specific surface area. The specific surface 
areas, as determined by BET nitrogen 
adsorption, of the two anomalous solids 
were 65 to 80 percent higher than the 
average of the remaining solids. To account 
for this variation, the available sulfide 
surface area per gram solid was estimated 
as the product of the sulfur content and 
the specific surface area. The correlation 
coefficient was 0.846 for the linear 
variation of sulfate release rate with 
respect to this available sulfide surface 
area. This dependence is consistent with a 
surface reaction dominating the sulfide 
oxidation . 

The r value increased to 0.962 when 
the data for the solid containing 1.17 
percent sulfur was omitted. Sulfate 
release from this solid was lower than 
predicted by. the linear model describing 
the remaining nine data points. This was 
the only sample containing calcium carbon- 
ate, which may have been responsible for 
the elevated surface area of this sample. 
The elevated pH of the drainage from this 
sample may also have inhibited the rate of 
sulfide oxidation (Nelson 1978). 

Some of the acid generated was con- 
sumed, or neutralized, by the dissolution 
of silicate minerals. Plagioclase, olivine, 
and pyroxenes are the silicate minerals 
which comprise the majority of the host 
rock. The dissolution of anorthite, which 
is a member of the plagioclase solution 
series, is represented by the following 
reaction (Stumm and Morgan 1981). 

CaAl 2 Si 2 O g (s) + 2H + (aq) + H 2 = 



2 + 



Ca" (aq) + Al 2 Si 2 5 (OH) 4 (s) 



(2) 



The dissolution of these minerals in 
laboratory experiments has often been 
reported as parabolic with respect to time. 
Initially the hydrogen ions in solution are 



184 



600 



500 - 



400 - 



300 - 



200 - 



1 00 - 




0.5 1.0 1.5 2.0 2.5 
PERCENT SULFUR IN SOLID PHASE 



illustrated semiquantitatively by comparing 
sulfate concentrations with the sum of 
calcium and magnesium values. The release 
of one mole of sulfate represents the 
production of two moles of acid, while the 
release of one mole of calcium or magnesium 
represents the consumption of two moles of 
acid (reactions 1 and 2, respectively). 
Initially the sum of calcium and magnesium 
concentrations exceeded the sulfate concen- 
tration, thus the ratio of the sum to 
sulfate was greater than one. This 
indicates that the rate of acid consumption 
by silicate mineral dissolution initially 
exceeded the rate of acid production by 
iron sulfide oxidation. This is borne out 
by the fact that the drainage pH was 
initially in the neutral range (fig. 4). 

The rate of acid production gradually 
increased over the first 18 weeks, but the 
rate of acid consumption increased conco- 
mitantly and neutralized the acid produced. 
After 20 weeks the acid production in- 
creased dramatically and the silicate 
mineral buffering was overcome. At this 
time, the ratio of the sum of calcium and 
magnesium concentrations to sulfate concen- 
tration dropped below a value of one and pH 
decreased, ultimately reaching 4.5. 



Figure 3. — Sulfide oxidation rate, as 
represented by mean sulfate 
concentration, versus percent sulfur. 
Points at 1.35 and 1.87 percent S 
were excluded from the regression 
analysis which determined line shown, 



rapidly and reversibly exchanged with 
alkalai ions on the mineral surface 
(Holdren and Berner 1979, Busenberg and 
Clemency 1976). The rate of release sub- 
sequently decreases, eventually becoming 
linear with respect to time (White and 
Classen 1979, Busenberg and Clemency 
1976). The rate of dissolution during this 
phase is relatively slow (Siegel 1981, 
White and Classen 1979, Busenberg and 
Clemency 1976). 

Dissolution of the silicate minerals 
was rapid enough to neutralize the acid 
generated by iron sulfide oxidation in the 
solids containing 0.8 percent S or less. 
The pH of drainage from these solids 
remained above 6. For the solids 
containing at least 1.35 percent S, 
reflecting the higher content of iron 
sulfides, the acid production rapidly 
overwhelmed the buffering capacity. The 
drainage from the solids containing 0.92 
and 1.24 percent S was initially neutral 
but gradually decreased, indicating that 
the rate of iron sulfide oxidation 
ultimately exceeded the dissolution rate 
of the silicate minerals present. 

The balance of acid production by 
iron sulfide oxidation and acid consump- 
tion by silicate mineral dissolution is 




Figure 4. — Drainage from 0.92 percent S 
solids versus time. Concentrations 
are in micromoles per liter, and R is 
the ratio of the sum of Ca and Mg 
concentrations to the sulfate 
concentration. 



185 



The buffering supplied by the sili- 
cate minerals was fairly uniform among the 
solids, despite differences in silicate 
mineralogy (table 2). Apparently these 
differences had little effect on the 
overall rate of silicate mineral dissolu- 
tion. The sample containing 1.87 percent S 
did exhibit slightly greater buffering and 
calcium release, and its composition is 
being examined in greater detail. 

The uniformity in buffering among the 
silicate minerals is particularly import- 
ant in predicting drainage pH. The primary 
copper-nickel mineralization of the Duluth 
Complex occurs as disseminated sulfides in 
a troctolitic-gabbroic series of rocks 
which intrudes an anorthositic and felsic 
series of older rocks (Weiblen and Morey 
1975). About 90 percent of the host rock 
is composed of silicate minerals, pre- 
dominantly plagioclase, olivine, and 
pyroxenes (Stevenson et al. 1979). If the 
buffering supplied by these minerals is 
somewhat uniform, the drainage pH will be 
controlled by the rate of acid generation 
due to oxidation of iron sulfides. 
Pyrrhotite is the predominant form of 
these iron sulfides (Bonnichsen 1972). 
Determination of the oxidation rate based 
on sulfur content will allow prediction of 
the acid-producing potential of mining 
wastes based on their sulfur content. 
Investigation of a wider variety of rocks 
will lend further insight into the 
influence of the various silicate minerals 
on buffering, as well as other factors. 

In contrast to the general uniformity 
in buffering, the solid containing 1.17 
percent S, despite exceeding the sulfur 
content of one of the acid solids, pro- 
duced drainage pH values which were 
typically above pH 8. Mineralogical 
analysis of this solid indicated a calcium 
carbonate content of about 3 percent. 
Dissolution of this calcium carbonate was 
reflected by calcium concentrations in 
drainage from this solid which were three 
to five times those in drainage from all 
but one other solid. Due to its rapid 
dissolution, the small amount of calcium 
carbonate was able to provide more 
buffering than the silicate minerals, 
which typically comprised more than 80 
percent of the solids. This further 
emphasizes the slow kinetics of, and 
limited buffering by, dissolution of the 
silicate minerals. 

The effectiveness of the silicate 
mineral buffering may have been further 
limited by the physical factors in the 
reaction environment. Iron sulfides can 
be oxidized by the oxygen and water vapor 
present in the air, while silicate dis- 
solution due to the presence of oxygen and 
water vapor is negligible. The acid 
resulting from iron sulfide oxidation may 
contact only the silicate minerals in the 
immediate vicinity of the iron sulfides. 
Dissolution of these silicates can leave a 
less reactive solid phase such as 
kaolinite (reaction 2; Busenberg 1978), 



which may further inhibit the transport of 
acid to other mineral surfaces and further 
limit buffering. This inhibition of 
buffering reactions may contribute to the 
increased metal recovery associated with 
the "resting" of leach columns and dumps in 
metallurgical work. 

In a submerged reaction environment, 
such as that in batch reactors, the acid 
generated could be readily transported to 
silicate mineral surfaces via the aqueous 
medium. As a result, batch reactor data 
may overestimate the buffering available in 
an environment subject to wet-dry cycling, 
such as stockpiles and unsubmerged mine 
walls. 

As drainage pH decreased, trace metal 
release increased, and this increase was 
quantified by linear regression analysis of 
the base 10 logarithm of metal concen- 
tration versus pH. For the log copper 
concentration versus pH plot, the slope was 
-1.11 (r=-0.944, n=38), and the corres- 
ponding slope for nickel was -0.99 
(r=-0.902, n=40). Similar trends were 
observed for cobalt and zinc. This in- 
crease in trace metal concentrations 
further magnifies the problems of acid 
drainage. The slopes indicate that as pH 
decreases one unit , copper and nickel 
concentrations will increase by about an 
order of magnitude. 



Comparison of Laboratory and Field Data 

Laboratory data on the test stockpile 
solids suggested that the smaller particles 
had a marked influence on the stockpile 
drainage quality. Particles smaller than 
2.0 mm in diameter comprised about 12 
percent of the total stockpile mass. The 
weighted average sulfur content of these 
particles was 1.0 percent as compared with 
0.6 percent S for the bulk rock. The sulfur 
content of these particles tended to 
increase as particle size decreased, with 
particles of diameter less than 0.053 mm 
having a sulfur content of 1.86 percent 
(Lapakko et al. 1986). 

The pH range generated in the labora- 
tory was quite close to that observed for 
the test stockpile drainage in 1982, when 
the solids were collected (fig. 5). 
Comparison with field data from subsequent 
years is somewhat tenuous since the stock- 
pile was dismantled, and the rock was 
divided among six different test plots. 
The diameter of the particles used in the 
laboratory tests was in the range of 0.053 
to 0.147 mm, and particles in this size 
range comprised only about 3 percent of 
the stockpile mass. The sulfur content 
of the particles was 1.26 percent, as 
compared with 0.6 percent for the bulk 
rock. The laboratory pH was slightly lower 
than that observed in the field. This 
suggests that the particle size range 
dominating the stockpile drainage chemistry 
may be slightly larger than that used in 



186 



X 

a 

Ui 
(3 

< 



6 - 



5 - 



T 



I 



1983 
1982 1984 



T 



x 



1 985 



|*— FIELD CONDITIONS 



Figure 5. — Drainage pH in laboratory 
compared to field values. Note: 
In September 1982 the test pile from 
which these laboratory solids were 
taken was dismantled and placed into 
test plots. Field data subsequent to 
1982 are from these plots. 



the laboratory, assuming similar sulfide 
oxidation rates in the two settings. The 
field rates may have differed from labora- 
tory rates due to fluctuations in ambient 
temperature and humidity, or inhibited by 
restricted oxygen transport. If so, 
another size range may have been the 
dominant influence on the stockpile 
drainage chemistry. Nonetheless, drainage 
pH from the 0.053 to 0.147 mm particles in 
the laboratory was quite close to that 
from the test stockpile in the field. 

Particle size is important for 
several reasons. As previously mentioned, 
sulfur content increased as particle size 
decreased. Sulfide minerals are softer 
than the silicate minerals and, therefore, 
are more prone to breakdown. Consequent- 
ly, they tend to concentrate in the 
smaller size fractions. Secondly, the 
surface area of the fines is relatively 
high since specific surface area increases 
as particle size decreases. Since the 
rate of iron sulfide oxidation was propor- 
tional to the available surface area, the 
rate of sulfide oxidation per unit mass of 
these particles is magnified. 

Thirdly, sulfide minerals in the 
small size fractions are most likely to be 
free rather than occurring as interstitial 
or included minerals. For Duluth Complex 
samples, the diameter at which sulfides 
are liberated from the host rock has been 
estimated as 0.074 mm (Weiblen and Morey 
1976); mineral benef iciation studies 
suggest a value of about 0.2 mm (Vifian 
and Iwasaki 1968). In particles larger 
than this critical size, a large fraction 
of the total sulfide surface area would be 
protected by the host rock, and therefore 
unavailable for reaction. Leaching tests 
run with larger particles tended to yield 
lower rates of sulfide oxidation and acid 
production. For example, the leaching of 



minus 10 mesh (diameter less than 2 mm) 
tailings containing 0.92 percent S produced 
neutral leachate (Natarajan and Iwasaki 
1983), while 0.053-mm to 0.147-mm solids of 
the same sulfur content produced acidic 
drainage in the present study. As dis- 
cussed earlier, the submerged reaction 
environment in the shake flask leaching of 
tailings may also have increased buffering 
efficiency by permitting acid contact with 
a larger silicate surface area. 

The pH of drainage from test stock- 
piles also decreased as the sulfur content 
increased. The most recent annual median pH 
values from piles containing 0.6 percent S 
were in the range of 5.25 to 6.5. The 
corresponding pH values for stockpiles with 
sulfur contents of 0.8 and 1.4 percent were 
about 4.25 and 3.6, respectively. These 
values suggest that stockpiles containing 
0.6 percent S are near the boundary of acid 
production. In the laboratory this boundary 
occurred between 0.80 and 0.92 percent S. 
This indicates that the laboratory tests 
simulate the leaching behavior of stock- 
piled solids having a bulk sulfur content 
roughly 68 to 79 percent of the laboratory 
solids. This "effective sulfur content" of 
stockpiled rock will, however, vary depend- 
ing on the particle size distribution of 
the stockpiled solids. Additional study on 
the effects of both composition and 
particle size on drainage quality are 
required in order to more accurately 
extrapolate laboratory results to stock- 
piles in the field. 

The laboratory data also closely 
simulated other aspects of stockpile 
drainage chemistry, indicating that the two 
reaction regimes were similar. The varia- 
tion of pH over time was similar in both 
cases. The drainage pH in the field tended 
to decrease over time, reflecting the 
decrease in buffering by silicate mineral 
dissolution. For the stockpiles containing 
0.6 percent S the pH decrease over time was 
gradual, but for the higher sulfur piles 
the decrease was relatively rapid (fig. 6). 
After the initial pH decrease, the drainage 
pH from the high sulfur piles was fairly 
stable. In the laboratory tests, a rapid 
pH decrease and subsequent period of 
relatively stable pH was also observed for 
samples with relatively high sulfur con- 
tents. The ratio of the sum of calcium and 
magnesium release to sulfate release traced 
pH fluctuations in the field as in the 
laboratory . 

Metals concentrations increased as pH 
decreased in both laboratory and field 
settings. The slopes for the log metal 
versus pH relationships, however, were 
steeper in the laboratory than in the 
field. As previously mentioned, the labora- 
tory data for the 0.92 percent solids 
indicated slopes of -1.11 for copper and 
-0.99 for nickel. Using the annual median 
values for metal concentration and pH for 
all stockpile drainages, the slopes for 
copper and nickel were -0.76 and -0.66, 
respectively (Eger and Lapakko 1985). 



187 



7 - 



6 - 



5 - 



4 - 




YEAR 



Figure 6. — Annual median pH values of test 
stockpile drainage from 1978 to 1986 
(modified from Eger and Lapakko 1985) 



The pH decrease and concomitant trace 
metal concentration over time were ob- 
served in drainage from an operational 
scale stockpile at the LTV Dunka site near 
Babbitt, MN. A drainage denoted as Seep 1 
flows from 1.5 million tons of stockpiled 
Duluth Complex rock removed from an open 
pit taconite mine. The stockpile was 
constructed in 1976 and drainage has been 
monitored since this time. From 1976 to 
1986 the annual median pH decreased from 
7.3 to 5.4 and annual median trace metal 
concentrations increased by factors of 
roughly 5 to 20 (Minnesota Department of 
Natural Resources 1988). 



content, consistent with first order depen- 
dence on available sulfide surface area. 
The rate of acid neutralization, which was 
controlled by silicate mineral dissolution, 
was relatively constant among the samples 
despite variations in the relative amounts 
of plagioclase, olivine, and pyroxenes 
present. This uniformity in neutralization 
rates adds credence to the possibility of 
predicting acid producing potential based 
on the sulfur content of Duluth Complex 
mining wastes, since they will be comprised 
primarily of these silicate minerals. At 
the 0.92 percent sulfur content, the rate 
of acid production due to iron sulfide 
oxidation exceeded the dissolution rate of 
silicate minerals which comprised 80 to 92 
percent of the solid phase. 

A solid containing 1.17 percent S did 
not produce acidic drainage since it 
contained about 3 percent calcium carbon- 
ate, or calcite, which is rare in the 
Duluth Complex (Stevenson et al . 1979). 
Unlike the silicate minerals, the calcite 
was able to neutralize the acid produced by 
the iron sulfide oxidation. The influence 
of this small mineralogical anomaly on 
drainage quality emphasizes the importance 
of examining the leaching behavior of 
numerous samples. Only then can leaching be 
described in terms of more general char- 
acteristics, such as sulfur content. 

The laboratory leaching results were 
consistent with field observations on the 
variation of pH with respect to percent 
sulfur and time. The range of 0.8 to 0.92 
percent sulfur in the laboratory appeared 
to be similar to a bulk sulfur content of 
approximately 0.6 percent for stockpiled 
rock. The drainage chemistry of stockpiles 
appears to be strongly influenced by the 
small particle size fraction, in which the 
softer sulfide minerals are concentrated. 
In conjunction with an elevated specific 
surface area, this produces a high sulfide 
surface area available for chemical 
reaction . 



CONCLUSIONS 

Ten Duluth Complex samples, obtained 
from core drilled by MINNAMAX exploration, 
were leached in a laboratory experiment. 
The number of samples was extremely small 
relative to the vast extent of the Duluth 
Complex. This experiment represents the 
starting point for examination of a wider 
compositional range, as well as the 
influence of other variables on the 
dissolution reactions. 



The equi 
drainage pH i 
mental time f 
state to be r 
tions, the dr 
stockpiles co 
decreased for 
equilibrium h 
Drainage pH f 
percent S app 
after two yea 
required for 



libration period for the 
s variable, and the experi- 
rame must allow for steady 
eached. Under field condi- 
ainage pH from three gabbro 
ntaining 0.6 percent S has 

nine years, indicating that 
as not yet been reached, 
rom a stockpile containing 0.8 
roached apparent equilibrium 
rs , while about one year was 
1.4 percent S rock. 



The su 
examined ra 
the variati 
content of 
pyrrhotite . 
of 0.8 perc 
drainage wh 
containing 
duced acidi 
sulfide oxi 
production , 



lfur content of 
nged from 0.47 
on due largely 
iron sulfides, 

Solids with s 
ent or less pro 
ile, with one e 
at least . 92 p 
c drainage. The 
dation, and the 
increased line 



the samples 
to 2 . 17 percent , 
to variable 
in the form of 
ulfur contents 
duced neutral 
xception, solids 
ercent S pro- 
rate of iron 
attendant acid 
arly with sulfur 



Although t 
stockpiles was 
observed in the 
concentrations 
Projections bas 
drainage from t 
underestimated 
and trace metal 
magnitude (Eger 
decrease in pH 
trace metal con 



he pH dependence for test 
not as strong as that 

laboratory, trace metal 
increased as pH decreased, 
ed on the first year of 
hese stockpiles would have 
subsequent release of acid 
s by at least two orders of 

and Lapakko 1985). The 
and attendant increase in 
centrations was also 



188 



observed in drainage from operational 
scale stockpiles. This temporal pH 
decline and trace metal concentration 
increase underlines the importance of 
allowing adequate reaction time for steady 
state to be approached. 



Acknowledgements 



The laboratory exp 
conducted by David Anto 
diligence and perceptiv 
experimental efficiency 
Albert Klaysmat , with a 
Drotts, was responsible 
precise chemical analys 
and numerous drainage s 
Paul Pojar and Milan Di 
to the construction of 
apparatus. Permission 
drill core samples was 
Kennecott Minerals Comp 



eriments were 

nson, whose 

eness improved the 

and accuracy, 
ssistance from Jean 

for the prompt and 
is of the solids 
amples generated, 
cklich contributed 
the experimental 
for use of the 
granted by 
any. 



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Kinetics and product of ferrous iron 
oxygenation in aqueous systems. 
Environ. Sci. Technol. 14: 561-568. 

Thingvold, D., Eger , P., Hewett , M. J., 
Honetschlager , B., Lapakko, K. , and 
Mustalish, R. 1979. Water Resources. 
In Minnesota Environmental Quality 
Board Regional Copper-Nickel Study 3 
(4): 217 pp. 

United States Bureau of Mines. 1980. A 
simple low-cost method for the 
dissolution of metal and mineral 
samples in plastic pressure vessels. 
BuMines RI 8480, 14 pp. 

United States Environmental Protection 

Agency. 1976. Quality criteria for 
water. U.S. Environmental Protection 
Agency, Washington DC, 256 pp. 

Veith, D. L. 1978. Minnesota copper- 
nickel resource processing model. 
Report by Regional Copper-Nickel 
Study. Minnesota Environmental 
Quality Board. May, 1978, 64 pp. 

Vifian, A., and Iwasaki, I. 1968. 
Mineralogical and benefication 
studies of the copper-nickel bearing 
Duluth gabbro. Trans. Soc. Mining 
Engineers, AIME 241: 421-431. 



Weiblen, P. W. , an 
Textural and 
characteristi 
the basal con 
Kawishiwi int 
Northeastern 
37th Annual M 
Annual Meetin 
AIME, Minneap 
25 pp. 



d Morey, G. B. 1976. 
compositional 
cs of sulfide ores from 
tact zone of the South 
rusion, Duluth Complex, 
Minnesota. I_n Proc. 
ining Symposium, 49th 
g, Minnesota Section, 
olis, MN, Jan. 14-15, 



Weiblen, P. W. , and Morey, G. B. 1975. 

The Duluth complex - A petrologic and 
tectonic summary. I_n Proc. 36th 
Annual Minnesota Mining Symposium: 
Dept. of Conferences and Continuing 
Education, Univ. Minnesota, 
Minneapolis, MN, pp. 72-95. 

White, A. F. , Classen, H. C. 1979. Dissolu- 
tion of silicate rocks-application 
to solute modeling. I_n Chemical 
modeling in aqueous systems, E. A. 
Jenne (ed.), American Chemical 
Society, Washington, DC, pp. 447-473. 



190 



ACID MINE DRAINAGE AND MANAGEMENT OF EASTERN OIL SHALE 
RESOURCES AND ASSOCIATED WASTE PRODUCTS 

Patrick J. Sullivan and Jennifer L. Yelton 1 



Abstract. --The objectives of this research were to 1) 
understand the chemistry of acid mine drainage associated with 
oil shale resources and 2) suggest appropriate management 
options to minimize water quality problems. In this study, two 
eastern oil shales were collected. Each shale was retorted and 
combusted to produce waste products representative of potential 
mining and energy conversion processes. Each raw, retorted, 
and combusted shale was studied under the following laboratory 
conditions; 1) oxidizing equilibrium, 2) oxidizing 
nonequilibrium, and 3) reducing equilibrium. The Experimental 
results show that 1) the fundamental chemical equations that 
have been used to predict acid mine drainage are not correct, 
2) using the new equations developed in this study, fundamental 
thermodynamic constants can be used to predict major and minor 
element activities, 3) the acid-base account method is not 
adequate for predicting acid potential, 4) simulated weathering 
methods should be used to predict acid potential, 5) combusted 
shales do not produce acid drainage under all disposal 
conditions, and 6) mineral weathering prior to the disposal of 
raw shales will significantly influence the acid potential of a 
material. These results suggest the following management 
implications 1) it is possible to predict leachate chemistry 
from waste and site specific data using thermodynamic 
constants, 2) placing weathered raw shale in a saturated 
environment can result in acid drainage, and 3) shales with a 
negative acid-base account should be combusted to minimize acid 
potential . 



INTRODUCTION 

When geologic materials containing iron 
sulfides are exposed at the earth's surface, acid 
mine drainage may occur. Because many raw and 
processed oil shales contain iron sulfides suggests 
that these materials need to be characterized to 
determine their acid potential prior to mining or 
processing. This characterization is a critical 
step in selecting management methods for a specific 
waste and disposal environment to avoid water 
quality problems associated with low pH and high 
trace element concentrations (Martin 1974, Griffin 
1980, Wahler 1978). 

The extraction of eastern oil shale resources 
is a critical concern, since these shales contain 
Iron sulfides. In addition to the raw shales, 



Patrick J. Sullivan is Manager of Waste 
Characterization and Chemistry, Western Research 
Institute, Laramie, WY. Jennifer L. Yelton 1s a 
Technician, Geotechnical Services, Western Research 
Institute, Laramie, WY. 



retorted and combusted oil shales may also contain 
iron sulfides (Padorek et al. 1984). Each of 
these materials will generate a leachate that 
reflects the mineral properties of each shale and 
selected disposal environment. The ability to 
characterize both acid potential and trace element 
release of each shale in a selected environment 
begins with understanding the chemical reactions 
that are commonly used to describe Iron sulfide 
oxidation. 

Theory of Acid Generation 

In a natural aerobic environment, Iron sulfide 
compounds have the potential to oxidize and produce 
acidity. The best example of this natural process 
1s the oxidation of FeS2 (pyrlte) that usually 
occurs in mining wastes, coal cleaning waste, 
spoil, and acid sulfate soils (Smith et al. 
1974). The oxidation of pyrlte In a natural 
environment 1s suggested by the following reactions 
(Stumm and Morgan 1981). 



Fe 



FeS 2 + 7/20 2 + H 2 

Fe + + l/40 2 + H + = Fe 3+ + 1/2 H 2 



+ 2SO4 

3j 



+ 2H + + 



(1) 



191 



Fe + + 3H 2 = Fe(0H) 3 + 3H + (3) 

3 2 2 

FeS 2 + 14Fe + + 8H 2 = 15Fe + + 2S0* " + 16H + (4) 

After pyrite is oxidized by atmospheric oxygen (eq 
1), ferric iron is produced extremely slowly (eq 
2). This second reaction, however, can be 
accelerated by microbial catalysis (aerobic) to 
increase the overall rate of ferrous iron 
oxidation. With the generation of ferric iron, 
insoluble ferric hydroxide may form (eq 3). Pyrite 
can also be oxidized by soluble ferric iron (i.e., 
in saturated or unsaturated environments), 
resulting in the generation of more acidity and 
ferrous iron (eq 4). 

With iron sulfide oxidation, Nordstrom (1982) 
proposes the formation of secondary iron phases 
that include ferrous iron sulfates, ferric iron 
sulfates (jarosite), and other ferric iron 
hydroxides and oxides. However, Nordstrom (1982) 
does not quantitatively show the environmental 
conditions in which these specific solid phases may 
be generated. Once these compounds are formed, 
Sullivan and Sobek (1982) have shown that soluble 
acidity in the form of ferric and ferrous iron will 
be released to solution. This is especially 
critical in evaluating the influence of geochemical 
weathering of iron sulfides prior to disposal. 

Research Objective 

Based on the discussion above, the objectives 
of this two-year study were to characterize the 
acid production of eastern oil shale waste products 
(i.e., raw shale fines, retorted shale, and 
combusted shale) as a function of process 
conditions, waste properties, and disposal 
practice and to recommend management practices. In 
order to understand acid production, each waste was 
characterized by determining the acid-base account 
and laboratory weathering studies to assess its 
acid potential and chemistry. 

DESCRIPTION OF RESEARCH 

Oil Shale Selection 

A high pyrite shale, selected from the U.S. 
Department of Energy's reference shale E86, was 
collected by Terra Tek. The sample was collected 
from an unweathered section of the upper part of 
the New Albany Shale of the Borden Formation 
located in Bullitt County, KY. A shale with a low 
concentration of pyrite was collected from a 
weathered outcrop of the Chattanooga Shale located 
in the Chestnut Mound section of Smith County, TN. 

Oil Shale Processing 

Both oil shales were crushed to pass a 3-inch 
screen. All particles passing a 1/2-inch screen 
were rejected from the feedstock. Each feedstock 
was 1) retorted using the indirect-heated Paraho 
process by Western Research Institute and 2) 
combusted using a fluidized bed process by J&A 
Associates. When both shales were retorted by the 
Paraho process and were indirectly heated to 500°C 
for 12 minutes, the spent New Albany Shale retained 
80% sulfur and had 28% organic carbon conversion. 
The spent Chattanooga Shale retained 76% sulfur and 
had a 26% organic carbon conversion when these 
shales were also combusted using a fluidized bed 
reactor in air at 760°C for 15 minutes, the 



combusted New Albany Shale retained 43% of the 
sulfur and had an 87% organic carbon conversion. 
The combusted Chattanooga Shale retained 81% of the 
sulfur and had an 83% organic carbon conversion. 
The details of the retorting and combustion 
activities are given in Sullivan et al. (1987). 

Laboratory Characterization 

The acid-base account was determined for each 
shale using the method as given by Sobek et al . 
(1978). Laboratory studies were designed to 
simulate the following disposal environments, 1) 
oxidizing nonequilibrium leaching, 2) oxidizing 
equilibrium, and 3) reducing equilibrium. All 
solutions generated from these laboratory studies 
were analyzed for the following: pH, Eh (as mv), 
cations by ICP (Al, Ba, Ca, Cd, Co, Cu, Na, Fe, K, 
Mg, Mn, Ni , Pb, Sr, Zn) and anions by ICP and IC 
(As, CI, Mo, S, Sb, Se, Si). 

Oxidizing Nonequilibrium Studies 

The humidity cell, developed by Caruccio 
(1968), can be used to simulate natural weathering 
of iron sulfides in an aerobic unsaturated 
environment. In this method, acid production will 
occur due to sequential leaching that allows 
oxidizing conditions to exist in a moist 
environment. In this study, the following humidity 
cell procedure was used. One kilogram of each 
shale (<2.00mm) was evenly spread out in the 
humidity cell, and the lid was tightly sealed. Dry 
air was passed over the samples for three days. 
From day four to day seven, humidified air was 
passed over the samples. On the seventh day, one 
liter of distil led-de ionized water was added to 
each cell and allowed to equilibrate for one 
hour. After that time period, the solutions were 
extracted, filtered, and the sediments were 
returned to each respective cell. This cycle 
continued for 19 weeks. A solution of Thiobacillus 
ferrooxidans and Thiobacillus thiooxidans was added 
to each cell on the fifteenth week to increase acid 
production. 

Oxidizing Equilibrium Studies 

This method simulates waste disposal in a 
saturated groundwater environment that will be in 
contact with atmospheric oxygen (i.e., fluctuating 
watertable). In this environment, it is 
anticipated that the reaction in equation 1 will 
occur. However, the chemistry of the system will 
be determined by thermodynamic equilibrium 
conditions (i.e., no leaching). 

Two hundred fifty-grams of each sample were 
placed into 500 mL Nalgene plastic bottles. Two 
hundred fifty milliliters of distil led-deionized 
water were added to these bottles and were capped 
tightly. A plastic tube was inserted through the 
cap to the bottom of the bottle to supply a 
constant flow of compressed air. Each sample 
bottle was placed into a water shaker bath at a 
constant temperature of 25°C. The samples were 
shaken at a constant rate of 60 rpm and were 
agitated manually four times a week. One bottle 
for each shale was removed from the water bath for 
analyses after 1, 4, 8, 16, 32, 64, 128, and 180 
days of equilibration time. The samples were 
filtered, and the clear filtrates were analyzed for 
pH, Eh, and total acidity and alkalinity and split 
for analyses. Samples saved for anion analysis 



192 



were kept at 4°C, and those saved for metals 
analysis were adjusted with concentrated HNO3 to a 
pH <2. 

Reducing Equilibrium Studies 

This method simulates waste disposal in 
groundwater conditions that would have limited 
contact with atmospheric oxygen. It would be 
expected that little or no oxidation of pyrite 
would occur from equation 1. However, it is 
anticipated that acid production could occur as a 
result of the reaction given in equation 4. There 
is no leaching so that chemical equilibrium will be 
established. 

Two hundred fifty-grams of each sample were 
placed into 500 ml Nalgene plastic bottles. Two 
hundred fifty milliliters of distilled-deionized 
water were added to these bottles and were capped 
tightly and taped to limit oxygen diffusion into 
the bottle. Each sample bottle was placed into a 
water shaker bath at a constant temperature of 
25 °C. 

The samples were shaken at a constant rate of 
60 rpm and were agitated manually four times a 
week. One bottle for each shale was removed from 
the water bath for analyses after 1, 4, 8, 16, 32, 
64, 128, and 180 days of equilibration time. The 
samples were filtered and the clear filtrates were 
analyzed for pH, Eh, and total acidity and 
alkalinity and split for analyses. Samples saved 
for anion analysis were kept at 4°C, and those 
saved for metals analysis were adjusted with 
concentrated HNO3 to a pH <2. 

RESULTS AND CONCLUSIONS 

Potential Acidity 

The results of the acid-base account are given 
in Table 1. These data show that the raw New 
Albany Shale, raw Chattanooga Shale and the 
retorted Chattanooga Shale have the potential to 
produce an acid mine drainage problem. Based on 
the work of Ferguson and Erickson (1986), it would 
also be predicted that these same materials would 
produce acidity in a field environment (i.e., 
acid-base account <+33). 

Table 1. — Acid-base account of raw and processed 
shales. 



Waste 


%FeS-S 


Acid 


Base 


Balance* 






(Tons 


of CaC0 3 /1000 


tons 








of waste) 






Raw NAS 


4.2 


-131.3 


4.9 




-126.4 


Retorted NAS 


0.08 


-2.5 


65.0 




62.5 


Combusted NAS 


0.04 


-1.3 


226.0 




224.7 


Raw CS 


1.9 


-58.4 


24.1 




-34.4 


Retorted CS 


0.8 


-25.0 


22.8 




-2.2 


Combusted CS 


0.6 


-18.8 


170.9 




152.1 



* (-) indicates a net acid potential 
NAS=New Albany Shale, CS=Chattanooga Shale 



Humidity Cell 

The humidity cell data represent a weathering 
environment that would stimulate maximum acid 
potential of a material. Humidity cell pH as a 
function of time is given in figure 1 for the New 
Albany Shale and the Chattanooga Shale in figure 
2. These data clearly show that the New Albany and 
raw and spent Chattanooga Shales will produce acid 
leachates. This suggests that these materials can 
produce acid mine drainage in an unsaturated 
surface disposal environment. 

When these data are compared to the acid-base 
account values of each shale, only the spent New 
Albany Shale was predicted to have a non-acid 
leachate (even in a field environment, i.e., value 
>+33). This suggests that the sources of 
neutralization in the spent shales were 
overestimated and/or not available for 
neutralization reactions during weathering. 

During initial leaching (first 28 days), the 
raw Chattanooga Shale shows an acid pH and 
decreases rapidly. The raw New Albany Shale, 
however, demonstrates an initial neutralization and 
a steady decrease in pH. Because the raw 
Chattanooga Shale was weathered (but a lower 
concentration of pyrite than the New Albany Shale), 
some soluble acidity was released to show an acid 
pH with little available alkalinity. 

Both retorted shales contain small 
concentrations of pyrite and both exhibit acid 
producing potential. After approximately 80 days, 
both shales establish a leachate pH below 4.0 after 
a loss of alkalinity. The combusted shales show a 
consistent decrease in pH with time. Like the 
retorted shales, the combusted shales have a small 
concentration of pyrite. However, because 
limestone was added to the combustion process, the 
combusted shales contain a higher neutralization 
potential. As a consequence, the leachate pH tends 
to remain higher. Even with this increased 
alkalinity, sulfide oxidation and dissolution of 
atmospheric carbon dioxide decreases leachate pH. 
When the pH was firmly established below 4.0 in the 
raw and retorted shales, microorganisms were added 
to the humidity cells. With this addition, the pH 
in both raw shales approached 2.0 (i.e., due to the 
increased rate of ferrous iron oxidation; equation 
2). A corresponding rapid decrease of leachate pH 
associated with the retorted shale was not as 
evident due to a much lower concentration of 
sulfides. 

Oxidizing Equilibrium 

In the simulated oxidizing groundwater 
environment, it is assumed that the flow rate of 
groundwater in contact with a specific waste 
material will be slow enough to establish chemical 
equilibrium. Since leaching does not occur in this 
system, pH will be controlled by secondary mineral 
formation and hydrolysis but not iron sulfide 
oxidation. The oxidizing equilibrium pH as a 
function of time for the New Albany Shale is given 
in figure 3 and for Chattanooga Shale in figure 4. 

The data for the raw shales show that the pH 
of the Chattanooga Shale extract is acidic and 
becomes more acidic with time, while the pH of the 
New Albany Shale extracts become alkaline and then 
pH decreases. 



193 



NEW ALBANY SHALE 



it- 



x 

Q. 



4- 



o° 

D °o°oo o 

OP 



o 



o 



a a D nao o^oo 



aa 



□ 



A 



o o o 
□a ° 

□□□□ DD 

'aa A aAaa Aa D AAQ 



40 60 SO 100 

HUMIDITY CELLS (time in days) 



—f— 

140 



Legend 

/S, RAW SHALE 

□ SPENT SHALE 

O COMBUSTED SHALE 



Figure 1.— pH data for the New Albany Shale humidity cell extracts. 

CHATTANOOGA SHALE 



it- 



»- 



•- 



x 

Q. 



• - 



4- 



t- 



o 



o 



o 
□ 



oo 



oo 



o o 



□ 



o 



o 



□ 



□ 






oo 



o 



aa a D nn o 

UDU nnn n n 

AAAAAAAAAAAgAgyffi 



-i r 

40 60 



T 



T 



— I — 

•0 100 120 

HUMIDITY CELLS (time in days) 



Legend 

A RAW SHALE 
□ SPENT SHALE 
O COMBUSTE SHALE 



Figure 2.~pH data for the Chattanooga Shale humidity cell extracts. 

194 



NEW ALBANY SHALE 



10- 



I 
o. 



^2 ° 

El A 



o 



Q 



-i 1 1 1 1 1 1 1 r - 

20 40 60 80 100 120 140 ISO ISO 



Legend 

^ RAW SHALE 

□ SPENT SHALE 

O COMBUSTED SHALE 



EQUILIBRATIONS (oxidizing doy) 



Figure 3.--pH data for the New Albany Shale oxidizing equilibrium extracts 

CHATTANOOGA SHALE 



14- 


o 










12- 


% 










K>- 


o 


o 


O 






8- 


D 










X 

a. 
s- 


tft 






o 


o 


4- 
2- 
0- 


^A 
A 

II- 


D 
A 


6 

i i — - 


A 
□ 

■| ■■ -,- T 1 


B 

— i 1 



Legend 

/\ RAW SHALE 

□ SPENT SHALE 

Q COMBUSTED SHALE 



20 40 60 80 100 120 140 ISO ISO 



EQUILIBRATIONS (oxidizing doy) 



Figure 4.— pH data for the Chattanooga Shale oxidizing equilibrium extracts. 



195 



This once again demonstrates the influence of 
mineral weathering prior to disposal (and not the 
concentration of pyrite). Both final pH values are 
still very low but not to the same degree as the 
humidity cell leachates. However, these data still 
support the fact that these raw shales will produce 
acid mine drainage if they occur in a similar 
groundwater environment. 

The retorted Chattanooga Shale also exhibits a 
pH trend that will eventually produce acid mine 
drainage; however, the New Albany Shale does not 
exhibit the same trend. Unlike the humidity cell, 
there is more time for neutralization reactions to 
occur. Since the retorted New Albany Shale has a 
greater neutralization potential than the retorted 
Chattanooga Shale, this trend is not unexpected. 
These data correspond more closely to the predicted 
acid potential as indicated by the acid-base 
account. 

The pH of all combusted shale solutions 
remains fairly alkaline to 180 days. Even with 
potential iron sulfide oxidation and recarbonation, 
the pH tends to remain high due to a greater 
concentration of alkaline materials (note: the com- 
busted New Albany Shale has a higher neutralization 
potential than the combusted Chattanooga shale). 

In this simulated environment, the acid-base 
account method more accurately predicts acid- 
producing potential. However, it still does not 



indicate the hazard of mineral weathering and 
soluble acidity. 

Reducing Equilibrium 

In the simulated reducing ground water 
environment, it is assumed that 1) the flow rate of 
groundwater in contact with a specific waste 
material will be slow enough to establish chemical 
equilibrium and 2) atmospheric oxygen will not 
Influence solution chemistry. Like the oxidizing 
equilibrium studies, leaching does not occur in 
this system, so pH will be controlled by secondary 
mineral formation and hydrolysis and not Iron 
sulfide oxidation. The reducing equilibrium pH as 
a function of time for the New Albany Shale is 
given in figure 5 and for Chattanooga Shale in 
figure 6. 

The data for the raw shales show that the pH 
of the New Albany Shale extracts remains just 
slightly acidic, while the Chattanooga Shale 
extracts are very acidic. This trend, once again, 
shows how soluble acidity can influence solution 
chemistry without pyrite oxidation. In the absence 
of 1) an atmospheric influence on solution 
chemistry and 2) leaching, retorted and combusted 
shale solutions does not generate acid mine 
drainage. 



x 
a. 



dP 

A 



NEW ALBANY SHALE 



o o 



□ □ 

A 
A 



O 







O 



El 



O 



-i— 

60 



-T - 
SO 



- 1 — 
100 



—f— 

120 



"I — 

140 



—r- 

160 



ISO 



Legend 

A RAW SHALE 

□ SPENT SHALE 

O COMBUSTED SHALE 



EQUILIBRATIONS ( reducing day) 



Figure 5.— pH data for the New Albany Shale reducing equilibrium extracts. 



196 



CHATTANOOGA SHALE 



X 

Q. 



o 



o o 



o 



o 



o 






D 



A A 



□ 



A 



□ 
A 



□ 
A 



20 



40 



"T - 

60 



100 



— r~~ 

leo 



— J— 

140 



—f— 
160 



Legend 

/\ RAW SHALE 

□ SPENT SHALE 

O COMBUSTED SHALE 



EQUILIBRATIONS (reducing day) 



Figure 6.— pH data for the Chattanooga Shale reducing equilibrium extracts. 



Generation of Acidity 

All total chemical analyses from each study 
were coded and speciated with the WATEQFC 
geochemical code (Runnells and Lindberg 1981) to 
calculate ion activities. Based on this 
speciation, the minerals that control iron and 
aluminum solubility in acid mine drainage solutions 
were determined. The results of this analysis are 
given in Sullivan, et al . (1988). The equations 
that control iron and aluminum solubility are given 
in figure 7 (Sullivan et al . 1988). The results of 
this study demonstrate that below a pH of 6.00, the 
activities of Fe 3+ and Al 3+ are controlled by basic 
sulfate solid phases. This shows that equations 3 
is not valid below a pH of 6.00. 

In addition, there is a significant 
interaction between Al 3+ and S0i+ 2 " influencing acid 
generation. Thus, the equations describing acid 
production associated with pyrite oxidation below a 
pH of 6.00 (most acid mine drainage is well below 
pH 6.00) should include the following reaction^: 
In aqueous solutions below pH 6.00, A10H 2 , 
A1(0H)2 + , Al2(0H)2 1 * + , FeOH 2+ , Fe(0H) 2+ , and 



Fe 2 (0H) 2 



4 + 



are the predominant ionic species. 



This suggests that at chemical equilibrium, pH will 
be a function of Al 3 and Fe 3 hydrolysis. 

Predicting Aqueous Chemistry 

In a disposal environment, the rate of 
infiltration or ground water flow will influence 
the aqueous chemistry. If waterflow is slow enough 



to allow contact times necessary to establish 
equilibrium (i.e., rate of reaction faster than the 
flow rate through the system), the water quality 
characteristics will reflect the influence of 
secondary mineral formation. With increased 
leaching rates and rapid removal of reaction 
products, the kinetics of mineral 
dissolution-precipitation and adsorption will 
determine water quality characteristics. This may 
result in a nonequilibrium or metastable 
equilibrium condition influencing the aqueous 
chemistry. 

If the waste/water system is at chemical 
equilibrium and the aqueous chemistry can be 
defined by secondary mineral reactions, then any 
leachate that leaves the disposal environment can 
be predicted using fundamental thermodynamic 
constants. With these fundamental relationships 
established, waste-specific/site-specific test 
methods can be used to predict water quality as a 
function of disposal site design (i.e., design 
which influences waste composition and flow rate). 

If water quality predictions are not 
acceptable, then treatment options can be included 
in the test procedure. 

Manage Acid Mine Drainage 

It is clear from the weathering study data 
that the acid-base account should be used as the 
first step in determining acid potential. 



197 



ENVIRONMENTAL 
CONDITIONS 



Geonydrology 

irvUliv* flow* 



Unsaturated Soil and/or m_i.ii . . 
Near Surtace Water Table IOM«»n» 



m FeS 2 ♦ \ 2 ♦ H 2 , * F. 2 + + 2S0 4 2 " + 2H+ 



Percolation of Acid Mine Drainage ._.... _ . , , 

Throuoh Soil or Into Groundwater 'Oxidizing or Reducing I 




Through Soil or Into Groundwater 



141 FeS 2 + 14Fe 3+ + »»;" T~* '«•" + 2S0 4 2 " + 16H* 
J ' 



1151 Fe 2 * + H 2 ■ * FelOHl* ♦ H* 

iiei 3Fe 2 + ♦ 4H 2 .=__;Fa3lOH) 2 * + 4H + 



(pe+pH>11.0, pH<3.0) 
l«7Fe 2 * ♦ S0 4 2 " + 7H 2 0^=^FeS0 4 -7H 2 0, s 




(pe + pH < 11.0, pH > 3.0) 
1171 Fe 3 * + SO4" + H 2 . * Fe0HS0 4 ls i + H 



* If deep groundwater conditions exist, 
Fe 2+ will not be oxidized. (i.e. equation 

#2) 



Fe 3 * Precipitation (3) Fe 3 * ♦ 3H 2 . ■ cminu\ 3m . 3H* 
3 (pe + pH < 11.0, pH >6) 



High and Low 
Leaching 



Aluminum Silicates 



FeS 2 Oxidation 
Precipitation I 

3 . pH<6.0 -a|3 , + soJ . + HjQ 

'"^0 

Al 3 * + H 2 



«22l Al 3 * * H 2 = 

123) Al 3 * + 2H 2 : 

124) 2AI 3+ + 3H 2 : 



AlOHSOns, + H + 121) 

Fe 3 * * H2O S 

AI(0H) 3 , s , + 3H* I5l Fe 3+ t 2 H 2 0: 

2Fe 3 * + 2H 2 ! 



IFeOH 2 * ♦ H* 118) 

SFelOHlj" + 2H* (19) 
2Fe2(0H)J* * 2H* 1201 



: AIOH 2 * + H* 
_eAI(OH>2 + 2H* 
_SAI 2 (0H)2 + + 2H* 






Figure 7. —Acid generation equations from Sullivan, Reddy, and Yelton (1988). 



With the completion of these tests, all 
samples with an acid-base balance <+33 should be 
tested using a weathering method that best 
simulates the disposal environment. It is also 
clear that more rapid weathering methods for 
determining potential acidity need to be developed. 

These data suggest that raw shales which are 
allowed to weather and generate soluble acidity may 
produce acid conditions in a saturated system. 
This demonstrates the need to evaluate leachate 
chemistry prior to disposal if shales are allowed 
to weather. The most significant result of this 
study is that combusted shales will not produce an 
acid leachate. Thus, shales with an acid-base 
account balance <+33 could be combusted prior to 
disposal. This could significantly reduce water 
quality problems and eliminate long-term water 
treatment costs. 



LITERATURE CITED 

Caruccio, F. T. 1968. An evaluation of factors 
affecting acid mine drainage production and 
the groundwater interactions in selected areas 
of western Pennsylvania. IN: Proceeding, 
Second Symposium on Coal Mine Drainage 
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Monroeville, PA. pp. 107-152. 

Ferguson, K. D. and P. M. Erickson. 1986. An 
overview of methods to predict acid mine 



drainage. IN: Proceedings of the Acid Mine 
Drainage Workshop. Halifax, Nova Scotia, pp. 
116-145. 

Griffin, R.A. 1980. Chemical and biological 
characterization of leachates from coal solid 
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Geological Survey, Urbana, IL. 

Martin, J. F. 1974. Quality of effluents from 
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Louisville, KY, 1974, pp. 

Nordstrom, D.K. 1982. Aqueous pyrite oxidation 
and the consequent formation of secondary iron 
minerals. IN: Acid Sulfate weathering: 
Kittrick, J. A., D.S. Fanning, and L.R. Hossner 
(ed.). Soil Sci. Soc. 

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P.F. Kimble, H. M. Atkins, S. R. Masuda, and 
R. H. Kratzer. 1984. Investigation of acid 
drainage potential of eastern oil shale. U.S. 
Department of Energy, DOE/LC/10843-1690. 

Runnels, D.D., and R.D. Lindberg. 1981. 
Hydrogeochemical exploration for uranium 
deposits: Use of the computer model 
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Sobek. 1974. Mine soil potentials for soil 
and water analysis. EPA 670/2-74-070. U.S. 
Environmental Protection Agency, Washington, 
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Sobek, A. A., W. A. Schuller, J. A. Freeman, and R. 
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Agency. Washington, DC. 

Stumm, W. and J. J. Morgan. 1981. Aquatic 
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Sullivan, P. J., J. L. Yelton, and K. J. Reddy. 

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600/7-78-222, Cincinnati, OH. 



199 



EVALUATION OF THE TCLP METHOD FOR TWO MILL TAILINGS J 



John C. Franklin and Eric G. Zahl 2 



Abstract. — Initial results are reported on 
determining the applicability of the toxicity char- 
acteristic leach procedure (TCLP) in predicting the 
potential for heavy metal contamination associated 
with mining wastes. TCLP tests, variations of the 
TCLP tests, and baseline tests were run on tailings 
samples from two mills to determine the sensitivity 
of laboratory protocol to slight procedural errors 
and to determine if any inherent factors produced a 
fatal flaw in using the test to evaluate mine tail- 
ings. Results from tailings A showed that metal ion 
concentration increased with higher liquid-to-solid 
ratios, with an increased volume of HOAc, and with 
longer mixing times. Results from tailings B showed 
the same general trends, but varied significantly in 
sensitivity to variations. The average yield of the 
standard TCLP tests as a percentage of the total 
digestion assay is 20% for tailings A and 10% for 
tailings B. This is a significant difference between 
the two types of tailings. Evaluation of these 
initial results indicates that the sensitivity of the 
three parameters to laboratory errors is probably 
acceptable if laboratory procedures are followed with 
normal attention to detail. These results are only 
the initial phase of a systematic evaluation of the 
TCLP method. Future research should emphasize how 
well the TCLP test actually simulates mine waste 
contamination phenomena. In particular, studies 
should focus on the applicability of the extraction 
fluid, chemical and mineralogical effects in and 
around the disposal area, time, pH effects, and 
oxidation effects. 



INTRODUCTION 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and 
Reclamation and the U.S. Department of 
the Interior (Bureau of Mines and Office 
of Surface Mining Reclamation and 
Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

2 John C. Franklin is a physical 
scientist and Eric G. Zahl is a civil 
engineer for the Spokane Research Center, 
U.S. Bureau of Mines, Spokane, WA. 



Substanti 
voiced regardi 
resources caus 
and resulting 
health. This 
Environmental 
to initiate a 
program for th 
in turn has cr 
contamination 
(EPA-December 
Members of the 
the mining ind 



al public concern has been 
ng contamination of water 
ed by mining (non-coal) 
adverse effects on human 
concern has prompted the 
Protection Agency (EPA) 
regulatory development 
e mining industry, which 
eated a need to predict 
potential at mine sites 
1985 Report to Congress) . 
American Mining Congress, 
ustry's most prominent 



200 



collective organization, have stated that 
resolution of the mining waste problem is 
the mining industry's number one issue. 
Regulatory efforts under the Resource 
Conservation and Recovery Act (RCRA) 
initially utilized the EP toxicity test 
as a primary standard with which to eval- 
uate contamination potential (EPA-Office 
of Solid Waste 1987) . This test was 
selected for analyzing metal mine wastes 
since its use is standard practice in 
determining whether industrial wastes are 
categorized as hazardous and unacceptable 
waste for sanitary landfills. However, 
there have been many technical concerns 
about its applicability to mining wastes. 
These concerns have centered mainly 
around the assumptions inherent in the 
testing procedures; that is, certain 
assumptions cause the test to be a poor 
simulator of mine waste disposal condi- 
tions. For example, the preferential 
dissolution of lead by the acetic acid 
used in the test procedures skews tests 
for lead, a heavy metal of great impor- 
tance in mining wastes (Lead Industries 
Association Report 1984) . 

As a result, the U.S. Bureau of 
Mines' Spokane Research Center (SRC) is 
performing research to determine more 
effective methods of evaluating the 
potential for contamination due to metal/ 
nonmetal mining activity. As part of 
this effort, we are examining existing 
laboratory testing methods to determine 
what applicability they may have in 
assessing contamination potential. The 
first test now in the process of being 
reviewed is EPA's toxicity character- 
istics leach procedure (TCLP) (Federal 
Register, Volume 51, No. 9) . This method 
was selected since it is currently being 
considered as a replacement for the EP 
toxicity test. Also, EPA is actively 
seeking evaluation methods for its Mine 
Waste Regulatory Development Program and 
it has been suggested that the TCLP 
method could be utilized. 

This report examines the results of 
tests using the TCLP method for monitor- 
ing metal contamination from mill tail- 
ings from two different mills. Organics 
and volatiles were excluded from these 
experiments because they are of little 
concern in mine tailings. Eleven key 
elements were analyzed. Some modifi- 
cations were made to the standard TCLP 
method in order to streamline laboratory 
procedures because there are no organics 
or volatiles to consider. For all tests, 
the organics testing portion was deleted 
and the leachant was filtered by gravity 
with a qualitative grade filter rather 
than by pressure or under a vacuum. 
Other experimental variations to the 
standard TCLP method were made as dis- 
cussed below to evaluate the method. 

MATERIALS AND METHODS 

To determine what changes in metal 
concentrations would be seen in the TCLP 



test, the liquid-to-solid 
strength, and mixing time 
The standard TCLP test pr 
a liquid-to-solid ratio o 
of 5.7 mL of glacial HOAc 
made up to 1 L of extract 
ASTM type 2 water to obta 
0.10 and a rotation time 
The liquid-to-solid ratio 
using a standard sample o 
and varying the total vol 
from 0.25 to 2.00 L. The 
HOAc was varied in our te 
0.20, and the duration of 
to 25 hours. 



ratio, acid 

were varied, 
ocedures require 
f 20:1, a volume 

(acetic acid) 
ion fluid with 
in a molarity of 
of 18 hours. 

was varied by 
f 50 of solid 
ume of slurry 

molarity of 
sts from 0.05 to 

rotation from 1 



Total digestion analyses and modi- 
fied TCLP tests using distilled water as 
an extraction fluid were also performed 
and the results compared to TCLP results 
to provide an indication of the strin- 
gency of the procedure. Tailings from 
two different mills were evaluated in 
these experiments. The first mill pro- 
cessed silver ore primarily while lead 
and zinc were secondary ores (tailings 
A) . The second mill processed lead ore 
(tailings B) . 

The analyses were conducted on an AA 
spectrometer for aluminum (Al) , silver 
(Ag) , calcium (Ca) , cadmium (Cd) , copper 
(Cu) , iron (Fe) , potassium (K) , magnesium 
(Mg) , nickel (Ni) , lead (Pb) , and zinc 
(Zn) . The tailings were not tested for 
organics or volatiles because the project 
is concerned with metal contamination 
only. Results reported here include only 
Al, Cu, Fe, Ni, Pb, and Zn because con- 
centrations of the other elements were 
very low or the variations were insigni- 
ficant. TCLP test results are normally 
presented in milligrams per liter of 
slurry. However, results of most of 
these tests were converted to milligrams 
per gram of solid for comparing with 
digestion assays. This conversion is 
made by using the equation 

A = (B*C)/D (1) 

where A = mg of contaminant in each g of 
waste, 
B = mg/L of contaminant in solu- 
tion, 
C = total volume of solution, L, 
and D = total g of solid used. 

This reporting unit (mg/g) also provides 
a more conceptually clear method of des- 
cribing the contaminant release from a 
given amount of waste material. 



In the firs 
both the liquid- 
acid strength we 
as shown in tabl 
tailings were us 
volume of leacha 
variations in th 
The data for tai 
increase in meta 



RESULTS 

t set of experiments, 
to-solid ratio and the 
re varied for tailings A 
e 1. Fifty grams of 
ed in each test and the 
nt was varied to obtain 
e liquid-to-solid ratio, 
lings A show a general 
1 concentrations (1) as 



201 



Table 1. --Leaching of metal ions from tailings A with varying molarity of acetic 
acid and volume of leachant, mg ion/g of solid waste. 



Volume. L 



_AX 



_QL 



_££_ 



ML 



_£fcL 



_&D_ 



0.05 Molar 



0.25 

.50 

.75 

1.00 

1.25 

1.50 

1.75 

2 .QQ 

0.25 

.50 

.75 

1.00 

1.25 

1.50 

1.75 

2.00 

0.25 

.50 

.75 

1.00 

1.25 

1.50 

1.75 

2.00 

Results using 



0.029 
.037 
.040 
.044 
.044 
.047 
.051 
■ 053 



0.009 
.024 
.026 
.026 
.027 
.027 
.027 
.027 



0.645 

1.08 

1.50 

1.88 

1.76 

2.05 

2.43 

2.54 



0.003 
.004 
.005 
.005 
.005 
.005 
.005 
.005 



1.07 
1.49 
1.84 
2.13 
2.15 
2.21 
2.01 
1.92 



0.145 
.149 
.137 
.124 
.158 
.162 
.161 
• 160 



0.10 Molar 



0.041 
.050 
.057. 
.063- 
.064 
.065 
.064 
.067 



0.024 
.026 
.028. 
.029- 
.028 
.028 
.029 
.02 9 



0.90 

1.50 

1.94 

2.31 1 

2.63 

2.89 

3.07 

3.81 



0.003 
.003 
.003. 
.003- 
.003 
.004 
.004 
■ 004 



1.26 

1.73 

2.12. 

2.49 ] 

2.39 

1.86 

2.54 

2. 8Q 



0.164 
.160 
.176. 
.180 J 
.340 
.174 
.182 
.172 






0.20 M ola r 



0.052 
.055 
.058 
.062 
.064 
.067 
.066 
.070 



0.026 
.028 
.029 
.029 
.029 
.029 
.029 
.030 



1.98 
2.83 
3.47 
4.00 
4.48 
4.66 
4.84 
5.33 



0.004 
.004 
.004 
.004 
.004 
.003 
.003 
■ 003 



1.50 
2.21 
2.25 
2.69 
2.79 
2.75 
2.94 
3 .04 



0.177 
.171 
.165 
.170 
.183 
.173 
.173 
■ 166 



standard TCLP test parameters (0.10 molar, 1.0 slurry) 



Table 2. --Leaching of metal ions from 

tailings B with varying acetic acid 
molarity and volume of leachant, 
mg ion/g of solid waste. 

Volume. L Cjj ££ £b_ Zn 

0.05 Molar 

0.50 0.002 0.001 0.187 0.090 

1.00 008 .001 .528 .102 

1.50 010 .003 .480 .118 

2.00 013 iM5 t£12 -128 

0.10 Molar 

0.50 0.009 0.001 0.338 0.125. 

1.00 013 1 .010 1 .562 1 .129 J 

1.50 014 .211 .674 .138 

2.00 015 && iMl -135 

0.20 Molar 

0.50 0.013 0.674 0.360 0.145 

1.00 011 .630 .720 .141 

1.50 016 1.58 .780 .154 

2.00 .£16. l,_8_2 ,£9_6. .1 48 

Result using standard TCLP parameters 
(0.10 molar, 1.0 L of slurry). 



the volume of leachant was increased 
(increased liquid-to-solid ratio) and 
(2) as the molarity or acid strength was 
increased for all three molarity groups 
except for Ni, which remained constant 
throughout the test. 

For tailings B (table 2) , the para- 
meters were varied, but fewer liquid-to- 
solid ratio data were collected since the 
results from the first set of experiments 
indicated the variations in concentra- 
tions were small and continually increas- 
ing with liquid amounts so that the mea- 
surement intervals could be made larger. 



Aluminum and nickel are not reported here 
because concentrations were low and 
changes in values were insignificant. 
The same general trends in test results 
occurred in tailings B as in tailings A. 
Concentration of the metal ions increased 
with increasing volume (liquid-to-solid 
ratio) and with increasing molarity of 
acetic acid in the leachant. 

In the third set of experiments, the 
mixing time parameter was varied for 
leachant in contact with tailings A 
(tables 3-6) . The standard method calls 
for 18 hours of mixing by rotation at 30 
rotations per minute. Four test contain- 
ers were rotated and samples were taken 
from containers 1, 2, 3, and 4 at 1, 2, 
3, and 5-hour intervals, respectively, 
for 25 hours. Multiple containers were 
used to determine if the removal of ana- 
lytical samples during rotation influ- 
enced the results. Results showed that 
if the data are averaged for each ele- 
ment and container, there is good agree- 
ment among the four different interval 
times. This indicates that removal of 
analytical samples during rotation did 
not influence test results. It should 
be noted that leachate concentrations of 
Cu, and to a lesser extent Al and Ni, 
remained relatively constant for various 
durations of mixing, while those of the 
other ions (especially Fe) continued 
to increase as the duration of mixing 
increased. 

The concentrations of the various 
metals obtained using the modified TCLP 
were compared with the total available 



202 



metal concentrations in the tailings. table 7 for the six reported metals and 

Assays were performed by the U.S. Bureau represent the digestion assay concentra- 

of Mines, Albany Research Center. The tion for each element in every gram of 

results of the analyses are reported in tailings. 

Table 3. — Effect of mixing time on the amount of metal ions leached from tailings 
A, container 1, mg ions/g of solid waste. 



Hours 


Al 


Cu 


Fe 


Ni 


Pb 


Zn 


1 


0.0406 


0.0246 


0.248 


0.0028 


1.86 


0.132 


2 


.0584 


.0256 


.428 


.0030 


1.82 


.138 


3 


.0490 


.0264 


.618 


.0030 


1.82 


.146 


4 


.0478 


.0264 


.730 


.0030 


1.88 


.146 


5 


.0484 


.0266 


.898 


.0034 


1.94 


.150 


6 


.0490 


.0266 


1.11 


.0034 


2.08 


.152 


7 


.0558 


.0276 


1.28 


.0040 


2.08 


.154 


8 


.0568 


.0274 


1.46 


.0036 


2.22 


.156 


9 


.0556 


.0276 


1.59 


.0038 


2.28 


.160 


10 


.0508 


.0270 


1.73 


.0040 


2.18 


.164 


11 


.0538 


.0268 


2.09 


.0040 


2.36 


.164 


12 


.0546 


.0274 


2.24 


.0040 


2.40 


.172 


13 


.0520 


.0266 


2.51 


.0040 


2.46 


.174 


14 


.0514 


.0272 


3.10 


.0042 


2.54 


.170 


15 


.0512 


.0274 


3.10 


.0040 


2.60 


.172 


16 


.0500 


.0270 


3.10 


.0036 


2.56 


.172 


17 


.0574 


.0272 


3.74 


.0042 


2.74 


.178 


18 


.0556 


.0272 


3.89 


.0040 


2.60 


.180 


19 


.0588 


.0266 


4.11 


.0034 


2.68 


.182 


20 


.0602 


.0266 


4.51 


.0034 


2.72 


.184 


21 


.0590 


.0264 


4.95 


.0044 


2.76 


.182 


22 


.0684 


.0258 


5.53 


.0042 


2.80 


.188 


Ave. .. . 


.0538 


.0267 


2.41 


.0037 


2.34 


.164 



Table 4. — Effect of mixing time on the amount of metal ions leached from tailings 
A, container 2, mg ions/g of solid waste. 



Hours 


Al 


Cu 


Fe 


Ni 


Pb 


Zn 


2 


0.0440 


0.0256 


0.400 


0.0028 


1.86 


0.122 


4 


.0484 


.0262 


.736 


.0032 


1.88 


.148 


6 


.0484 


.0264 


1.05 


.0034 


1.98 


.146 


8 


.0574 


.0272 


1.59 


.0038 


2.16 


.164 


10 


.0504 


.0266 


1.83 


.0042 


2.28 


.158 


12 


.0536 


.0274 


2.11 


.0044 


2.32 


.168 


14 


.0504 


.0270 


2.40 


.0040 


2.44 


.168 


16 


.0500 


.0270 


2.74 


.0056 


2.46 


.172 


18 


.0512 


.0272 


3.10 


.0040 


2.42 


.176 


20 


.0530 


.0268 


3.38 


.0034 


2.60 


.178 


22 


.0440 


.0266 


3.64 


.0042 


2.58 


.176 


24 


.0530 


.0266 


3.84 


.0040 


2.56 


.180 


25 


.0544 


.0272 


4.03 


.0042 


2.65 


.182 


Ave 


.0506 


.0268 


2.37 


.0039 


2.32 


.164 



Table 5. — Effect of mixing time on the amount of metal ions leached from tailings 
A, container 3, mg ions/g of solid waste. 



Hours 


Al 


Cu 


Fe 


Ni 


Pb 


Zn 


3 


0.0488 


0.0264 


0.550 


0.0026 


1.82 


0.150 


6 


.0490 


.0268 


1.03 


.0034 


2.00 


.158 


9 


.0512 


.0268 


1.55 


.0046 


2.20 


.164 


12 


.0532 


.0272 


2.13 


.0040 


2.44 


.168 


15 


.0526 


.0274 


2.52 


.0032 


2.48 


.170 


18 


.0496 


.0274 


2.97 


.0040 


2.48 


.176 


21 


.0446 


.0258 


3.21 


.0036 


2.54 


.180 


24 


.0518 


.0270 


3.64 


.0038 


2.54 


.180 




.0544 


.0276 


3.75 


.0042 


2.62 


.182 


Aye. . . . 


.0506 


.0269 


2.37 


.0037 


2.35 


.170 



203 



Table 6. --Effect of mixing time on the amount of metal ions leached from tailings 
A, container 4, mg ions/g of solid waste. 



Hours 


Al 


Cu 


Fe 


Ni 


Pb 


Zn 




0.0492 


0.0258 


0.882 


0.0034 


1.88 


0.148 


10 


.0504 


.0258 


1.74 


.0032 


2.26 


.156 




.0496 


.0262 


2.44 


.0032 


2.32 


.160 


18 


.0488 


.0258 


2.89 


.0034 


2.36 


.170 


20 


.0550 


.0260 


3.03 


.0040 


2.44 


.170 


25 


.0514 


.0260 


3.45 


.0040 


2.38 


.172 


Ave. . . . 


.051 


.026 


2.41 


.004 


2.27 


.163 



Table 7. — Assay results for tailings A and B, mg/g, and standard TCLP results as 
a % of assay. 



T a ilings 



.Al. 



_£1L 



_£e_ 



_NL 



-EJ2- 



-ZjL 



Ave. 



A, mg/g (assay) 

% TCLP is of assay. 



B, mg/g (assay) 

% TCLP is of assay. 



0.199 0.141 68.1 0.015 5.41 1.60 

32 20 4 20 36 10 20 

.188 .318 27.7 .138 1.77 .934 
6 4 Q 4 3J 14 IQ_ 



Table 8. --Average of 10 standard TCLP tests for tailings A and B, mg/g and 
standard deviations (SD) , and coefficient of variation (CV) . 

Tailings Al Cjj £e Ni P_b. Zn 

A 0.064 0.028 2.99 0.003 1.97 0.167 

SD .001 .0004 .284 .0004 .036 .016 

CV 2 1 9 13 2 10 

B 0.012 0.012 0.011 0.005 0.524 0.133 

SD .003 .0004 .003 .0004 .132 .005 

CV 25 2 21 8 25. Q_ 

Table 9. --Metal concentrations using deionized water as leachant, ions/g of 
solid waste. 



Tailinas 


Al 


Cu 


Fe 


Ni 


Pb 


Zn 


A 


0.0002 
.0024 


nd 

nd 


nd 
0.0002 


nd 
nd 


0.0014 
.0018 


0.0050 
.0004 



nd - below detectable limits 



Standard TCLP test results are com- 
pared to the baseline assay data in table 
7. Note that tailings A had an average 
of 20% for the six metals while tailings 
B had an average of only 10%. 

Ten individual analyses were made on 
both tailings samples to determine the 
statistical reliability of each metal 
concentration (table 8) . The coefficient 
of variation was less than 15 for tail- 
ings A in most of the metal concentra- 
tions, while tailings B had three ele- 
ments with a coefficient of variation of 
25 or larger. However, concentrations of 
two of these three elements approached 
the detection limits of the spectrometer. 
There is good agreement between the 
18-hour test (tables 3-6) and the data 
reported in table 8. 

The leachability of these ions by 
rainwater may be considerably less than 
that inferred from TCLP results using 
0.1M. acetic acid as leachant. Results of 
a modified TCLP test in which 1 L of 
deionized water was mixed with 50-g 



samples of tailings A and B are reported 
in table 9. Significantly less amounts 
of each element were leached by the 
water. 

As previously stated, mill A was 
processing silver ore while mill B was 
processing lead. Because the ores from 
the two mills were different, the mill 
waste reacted quite differently to 
testing. The principal minerals present 
in mill A waste were silica, galena, and 
sphalerite; the principal mineral from 
mill B tailings was dolomite with trace 
amounts of sphalerite, pyrite, and 
galena. 

A major difference observed during 
our investigation of these two mill tail- 
ings samples was the change in pH during 
the mixing period. The TCLP requires a 
starting pH of 2.88 but, unlike the EP 
toxicity method, does not specify the 
ending pH. For all our tests, the pH was 
measured before adding the mill waste and 
after the 18-hour mixing period. The pH 
increased to about 3.7 for mill tailings 



204 



A and to 5.0 for mill tailings B. Such 
an increase would be expected. 

In another series of tests for 
tailings A, liquid samples (ready for AA 
analysis) were used to evaluate the time 
affect of sample storage. The samples 
were stored at room temperature and rerun 
after 2 weeks and again after 3 weeks to 
determine if delayed analysis affected 
the results. The average of all 52 runs 
was determined for Fe, Pb, and Zn and 
compared to the data in table 8. The 
results showed less than 6% difference 
between any of the delayed samples as 
compared to the 10 runs reported in table 
8, indicating little change because of 
the 3-week delay before the filtered 
samples were analyzed. 

CONCLUSIONS 

The current project is the first 
phase of a detailed evaluation of the 
TCLP test to determine what, if any, 
applicability the test has in evaluating 
mine wastes. An evaluation of the re- 
sults shows that there are many factors 
that may influence the TCLP laboratory 
test. The following points summarize the 
research to date. 



Based on these results and other 
similar research at SRC, it appears that 
better laboratory assessment methods 
could be developed that more appropriate- 
ly aid in the simulation or prediction 
of contamination from mine wastes. This 
preliminary assessment was made because 
of the great diversity in mineralogy, 
geology, hydrology, processing, and dis- 
posal methods of different mines, and 
because an accurate assessment of the 
hazards posed by any individual mining 
waste site is very difficult. Even in 
instances where the same material is 
mined, the disposal of waste may produce 
significantly different contamination 
effects from site to site. Significant 
factors, such as the effects of geologic 
formations and the hydrogeology between 
the source of the contaminant and the 
potential point of impact, should be con- 
sidered. Such an evaluation may require 
assessment methods other than laboratory 
tests, such as field geophysical surveys. 
It is critical that these factors be 
taken into account when developing an 
effective assessment method for evalu- 
ating potential contamination from mine 
wastes. 

LITERATURE CITED 



The TCLP results are not highly 
sensitive to changes in the liquid- 
to-solid ratio, acid strength, and 
duration of mixing. Small labora- 
tory measuring errors will not have 
a significant impact on TCLP test 
results unless gross errors are 
made. 



American Resources Corporation and 
Environmental Engineering and 
Management. 1984. Technical eval- 
uation of the U.S. EPA extraction 
procedure. Report to Lead Indus- 
tries Association and Cadmium 
Council. 26 pp. Lead Industries 
Association, Inc., New York, NY. 



Variations of the length of mixing 
time for the TCLP test show that 
equilibrium conditions for metal 
concentrations are attained for only 
some elements during the standard 
18 hours. Other metal concentra- 
tions were increased at varying 
rates, indicating that the TCLP test 
results may be arbitrarily biased 
for certain elements as influenced 
by the length of mixing time. 



The pe 

using 

f icant 

tested 

indica 

to bei 

alogic 

which 

consid 

test 



rcentage of 
the TCLP me 
ly between 

(20% versu 
te a bias o 
ng influenc 
al properti 
are not lea 
ered in the 



metals extracted 
thod varied signi- 
the two tailings 
s 10%) . This may 
f the TCLP method 
ed by the miner- 
es of the waste, 
ching phenomena 

standard TCLP 



Further research is needed to deter- 
mine the applicability of the TCLP 
test to mining waste. The scope of 
this initial effort was limited to 
two tailings sources and assessment 
of only a few factors that could 
influence the applicability of the 
TCLP test to mining wastes. 



Doepker, R. D. 1987. 
of factors influ 
lution of metals 
tailings. Proce 
Drainage and Sur 
tion Conference 
American Society 
and Reclamation 
ment of the Inte 
OSMRE) , Pittsbur 



The interrelation 
encing the disso- 

in columns of mine 
edings, 1988 Mine 
face Mine Reclama- 
sponsored by the 

for Surface Mining 
and the U.S. Depart- 
rior (USBM and 
gh, PA. 



Federal Register, Volume 51, No. 9. 

1986. Appendix I. Toxicity charac- 
teristic leaching procedure (TCLP) . 
January 14, pp. 1750-1758. 

U.S. Environmental Protection Agency, 
Office of Solid Waste. 1985. 
Wastes from the extraction and 
benef iciation of metallic ores, 
phosphate rock, asbestos, overburden 
from uranium mining, and oil shale 
- report to Congress. 300 pp. 

U.S. Environmental Protection Agency, 

Office of Solid Waste. 1987. RCRA 
subtitle D regulatory program for 
the management of mining wastes - 
draft regulatory development plan. 
180 pp. 



205 



A COMPARISON OF RESULTS FROM ACID-BASE ACCOUNTING VERSUS POTENTIAL 

ACIDITY MEASURED BY THE PEROXIDE OXIDATION OF WEATHERED AND 

UNWEATHERED SOILS CONTAINING PYRITE 1 



2 

Ammons and P. A. Shelton 



ABSTRACT 

Ammons, J. T. (1) and Shelton P. A. (2). ((1) 
University of Tennessee, Knoxville (2) Southern 
Illinois University at Carbondale). Fifty samples 
from weathered disposal area soils and forty-nine 
samples from unweathered overburden cores were anal- 
yzed using two laboratory methods to assess potential 
acidity. Total sulfur values (percent) were compared 
with potential acidity values by hydrogen peroxide 
oxidation ( meqH+/100g ) . The percent total sulfur 
values for the unweathered samples (cores) ranged 
from 0.006 to 1.32 percent. Potential acidity by 
hydrogen peroxide oxidation ranged from 3.05 to 67.98 
meqH+/100g. Total sulfur values for disposal area 
soil (weathered) ranged from 0.008 to 0.920 percent. 
Hydrogen peroxide oxidation values ranged from 0. to 
52.34 meqH+/100g in the weathered disposal area soils. 
The coefficient of determination (r 2 ) comparing the 
two methods is 0.89 for the unweathered samples and 
0.71 for the weathered samples. There is a positive 
correlation between the two methods for the unweather- 
ed (fresh) core samples. This relationship is less 
pronounced for weathered samples from the surface of 
the disposal areas. 



ipaper presented at the 1988 Mine Re- Southern Illinois University at Carbondale. 
clamation Conference Sponsored by the 
American Society for Surface Mining 
and Reclamation and the U.S. Depart- 
ment of the Interior (Bureau of Mines INTRODUCTION 
and Office of Surface Mining Reclama- 
tion and Enforcement), April 19-22, Laboratory procedures to assess poten- 
1988, Pittsburg, PA. tial accidity problems prior to mining or 
2 J. T. Ammons, Associate Professor, before major construction have been used 
University of Tennessee, Knoxville in overburdern analyses studies by many 
and P. A. Shelton, Research Associate, 

206 



investigators (West Virginia University 
1971, Smith et al . 1974; Smith et al. 1976; 
Ammons et al . 1983). These laboratory tech- 
niques allow for postmining landuse and 
post construction landuse planning and deci- 
sion-making . 



hist 
acid 
mini 
proc 
dete 
oxid 
this 
pare 
tota 
lect 
posa 
Tenn 
Coun 



Two 
oric 
-pro 
ng a 
edur 
rmin 
e ox 
stu 
the 
1 su 
ed f 
1 ar 
esse 

ty 



laborato 
ally used 
ducing ma 
nd constr 
es are ac 
ation of 
idation ( 
dy, sampl 
se proced 
lfur valu 
rom fresh 
eas along 
e Tombigb 
MS. 



ry p 

to 
teri 
ucti 
id-b 
pote 
Smit 
es w 
ures 
es. 
ove 
the 
ee W 



roce 

esti 

als 

on a 

ased 

ntia 

h et 

ere 

ove 

The 

rbur 

div 

ater 



dures 
mate t 
result 
ctivit 

accou 
1 acid 

al. 1 
analyz 
r a wi 

sampl 
den co 
ide se 
way in 



have been 
he levels of 
ing from 
ies. These 
nting and 
ity by per- 
974). In 
ed to com- 
de range of 
es were se- 
res and dis- 
ction of the 
Tishomingo 



In 1971/ West Virginia University com- 
pared these procedures using fresh over- 
burden rock samples from surface mines and 
found a high correlation between the two 
tests. The same comparison was conducted 
on old minesoil samples and no consistent 
relationship was found between the two tests 
(WVU 1971) . 

During the excavation of the divide 
section of the Tennessee Tombigbee Water- 
way pyritic spoil material was excavated 
and deposited in the disposal areas. The 
lower part of the Eutaw formation (Cretace- 
ous) contained lenses of pyrite concentrated 
in sandy nodules (Ammons et al. 1983). 

Corps of Engineers personnel used the 
peroxide oxidation procedure onsite to 
access potential acidity of waterway slope 
soils and requested a comparison of acid- 
base accounting (as estimated from total 
sulfur) and the peroxide oxidation proced- 
ure (WVU 1971). Although acid-base account- 
ing is less time consuming laboratory faci- 
lities are not always available to complete 
this procedure. This was the impetus to 
conduct a comparison of the two methods. 



will not allow the peroxide to decompose 
upon heating if no sample is present to 
catalyze the decomposition. This would re- 
sult in interference from the peroxide (a 
weak acid) during the titration. Addition- 
ally, the carbonates and sulfates were not 
removed from the samples prior to oxidation. 
Leaching the samples is a time-consuming 
process, so the hydrogen peroxide analyses 
were completed under the same conditions 
used by Corps of Engineers personnel in the 
field. 



RESULTS AND DISCUSSION 



The 
drogen pe 
acid-base 
CaCC>3 equ 
are shown 
ranged fr 
2 shows s 
.920 perc 
pies. Su 
percent o 
the inten 
at the su 



total 
r 100 

acco 
iv.) 

in t 
om 0. 
ulf ur 
ent o 
lfur 
n dis 
se ch 
rf ace 



sulfur 

grams 
unt (to 

for ov 
he tabl 
006 to 

values 
n selec 
values 
posal a 
emical 

immedi 



, millequ 
of materi 
ns/1000 t 
erburden 
e 1. Sul 
1.328 per 

ranging 
ted dispo 
rarely ex 
rea soils 
oxidizing 
aely afte 



ivalents hy- 
al, and final 
ons material 
core samples 
fur values 
cent. Table 
from .008 to 
sal area sam- 
ceeded 1.0 
because of 
environment 
r placement. 



tota 

drog 

la-b 

pies 

tion 

Figu 

ods 

rela 

sugg 

cedu 

la 



The 

1 sul 

en pe 

Th 

(fig 

(r2 = 

re la 

on we 

tions 

estin 

re f r 



resu 
fur 
roxi 
e re 
. lb 
0.89 
sho 
athe 
hip 
g in 
om s 



Its of the 
and potent 
de are ill 
suits of t 
) show a s 
) between 
ws the res 
red dispos 
is less pr 
terf erence 
ome weathe 



compariso 
ial acidit 
ustrated i 
he fresh c 
ignif icant 
the two me 
ults of th 
al area sa 
onounced ( 

in the pe 
ring produ 



n between 
y by hy- 
n figures 
ores sam- 
correla- 
thods . 
e two meth- 
mples. The 
r2=0.71), 
roxide pro- 
cts ( fig . 



Figures 2a-b present the results of 
the peroxide procedure compared with acid- 
base accounting. Acid-base accounting is 
reported in calcium carbonate equivalent 
(tons/1000 tons material) and includes neu- 
tralization potential (amount of bases 
present ) . 



MATERIALS AND METHODS 

Forty-nine samples from fresh over- 
burden cores ( unweathered) and fifty samples 
of disposal area soils (weathered) were 
selected for this study. Sulfur values 
ranged from 0.006 percent to 1.328 percent, 
representing the range of sulfur values re- 
ported the divided section of the waterway. 

Total sulfur analyses (as part of acid- 
base accounting) were completed on all sam- 
ples using an Automatic Sulfur Analyzer. 
All samples were ground to 60 mesh for lab- 
oratory analyses as described by Sobek et 
al. (1878). 

Potential acidity by hydrogen peroxide 
was completed according to Smith et al . 
(1974) except for the following modifica- 
tions. Blanks were not used because the 
stablizer present in the hydrogen peroxide 



207 



Tabla 1. --Total sulfur, aUllisquivalmtJ of hydre|»n 07 paroaida 
aaldstlot), and acld-a**a account of aaaplaa tram 
erarburdan coraa, Taftnaaaa* -TcaA) i a, aaa Natarvay. 



Sa« a la 


Percent 


Nee I*/ 


Acld-Uee 


Number 


Sulfur 


100 c 


Ac cow t 

CaCO. Equivalent 
Torn/ 1.000 Torn Malarial 








195-14 


1.3210 


67.91 


-45.12 


•954-11 


1.2100 


51.40 


-23.11 


•97-22 


1.2350 


43.33 


-57.75 


•95-23 


1.0620 


46.61 


-24.19 


•97-21 


1.0400 


35.96 


-43.04 


197-34 


0.1194 


4.97 


-21.25 


•97-J6 


0.1573 


16.36 


-16.01 


•95-26 


0.7(20 


33.45 


-21.37 


•94-30 


0.7495 


29. (2 


-24.65 


•94-5 


0.7101 


33.16 


-25.17 


•97-51 


0.6488 


30.62 


-30.13 


•94-33 


0.5532 


23.91 


-19.75 


•94-19 


0.5496 


(.•0 


-22.35 


•97-23 


0.S4I1 


24.47 


-23.14 


•97-41 


0.5391 


11.01 


-24.96 


•95-35 


0.5261 


22.73 


-19.67 


•94-6 


0.5090 


21.97 


-22. (1 


•97-20 


0.4903 


17.13 


-18.62 


•96-42 


0.4884 


20.31 


-15.95 


•95-31 


0.4632 


17.21 


-12.26 


•94-14 


0.4504 


20.48 


-14.06 


•94-3« 


0.4375 


20.05 


-17.12 


•94-34 


0.4136 


17.43 


-14.65 


•94-44 


0.4074 


14.64 


-14.45 


•94-3 


0.3904 


11.92 


-14.66 


•94-36 


0.3794 


15.98 


-13. St 


•96-34 


0.3470 


14.07 


• 1.32 


•97-13 


0.3333 


14.73 


-11.97 


•95-10 


0.307* 


12.95 


- 9.29 


•97-47 


0.2159 


•.24 


-10. (5 


•95A-12 


0.2749 


10.68 


- «.47 


•94-15 


0.2623 


9.54 


- 7.21 


•954-9 


0.2522 


11.14 


- 7.06 


•94-21 


0.2450 


1.1* 


• 6.11 


•95-21 


0.2350 


9.03 


- 5.99 


•96-40 


0.2046 


(.46 


- 5.47 


•97-42 


0.1915 


7.71 


- 7.04 


•97-12 


0.1776 


13.32 


- 7.76 


•95-10 


0.1566 


7.34 


- 5.05 


•97-10 


0.1333 


6.11 


- 5.39 


•954-5 


0.1330 


• .91 


- 4.32 


•94-12 


0.1131 


5.74 


- 2.54 


■9SA-I 


0.1046 


3.18 


- 2.20 


•9SA-4 


0.0917 


• ••3 


- 2.99 


•954-13 


0.0603 


4.17 


- 1.02 


•97-53 


0.0517 


3.05 


0.23 


•97-17 


0.0213 


7.34 


0.40 


•96-26 


0.0121 


4.62 


0.54 


•96-23 


0.0064 


3.14 


0.72 



labia— 2. Total au 


fur. ■llllaqul> 


alanta o: 


bydrosan by peroalde 


Olidatlo 


I, and acld-baal 


account 


of taaplai takan fro* 


ovarburdan coraa, Tannai 


aaa-ToBbl|baa watarvay. 


SaapLe 


Parcant 


Kaq 8*/ 


Acld-Baaa 


Number 


Sulfur 


100 C 


Account 
CaCo Equivalent 
Tona/1.000 Tons Hatorlal 








1503-02 


0.9203 


52.34 


-37.07 


501 -Nl 


0.5772 


0.00 


20.47 


1504-P1U 


0.S013 


19.23 


-22.59 


1S03-L3 


0.4926 


24.06 


-17.03 


1504-11 


0.4633 


17.71 


-26.94 


501-0415 


0.4554 


4.76 


-13.76 


1504-12 


0.4386 


16.91 


-23.40 


1203-02 


0.4362 


16.40 


-19.25 


501-J1 


0.4153 


0.00 


203.92 


1503-P1L5 


0.3786 


24.61 


-13.21 


1503-42 


0.3671 


17.28 


-19.31 


1504-P1L4 


0.3390 


16.17 


-11.03 


1704-P1L3 


0.3016 


12.58 


-14.12 


1503-F2L4 


0.2462 


9.28 


-11.84 


1204-L3 


0.2240 


9.98 


- 7.92 


1503-P1L4 


0.2218 


15.27 


- 1.31 


1204 -Ql 


0.2111 


0.82 


1.06 


1205-05174 


0.1863 


2.74 


-6.58 


1704-P1L1 


0.1629 


7.92 


-11.19 


1203-L2 


0.1582 


6.86 


-6.35 


1508-04834 


0.1410 


1.01 


10.76 


1503-05845 


0.1345 


0.00 


23.77 


1204-D2 


0.1098 


5.73 


-2.36 


1701-07173 


0.1091 


0.52 


14.45 


1504-03150 


0.0999 


6.47 


-3.12 


1503-F1 


0.0943 


4.13 


-2.03 


1203-72 


0.0920 


7.18 


-4.29 


1S04-H1 


0.0869 


4.76 


-5.95 


1504-82 


0.0803 


6.08 


-4.44 


1203-P1 


0.0797 


0.00 


24.27 


1203-111 


0.0722 


1.37 


-1.56 


1204-H1 


0.0664 


0.00 


21.11 


1203-01 


0.0633 


0.00 


16.56 


1204-11 


0.0597 


1.99 


0.57 


1204-42 


0.0565 


3.86 


-1.92 


1203-81 


0.0503 


0.00 


14.86 


1203-61 


0.0492 


0.00 


14.86 


1203-E1 


0.0474 


0.16 


2.74 


501-C1 


0.0411 


4.03 


-1.75 


501 -El 


0.0373 


0.00 


9.37 


501 -E2 


0.0350 


0.67 


-0.66 


501 -13 


0.0337 


1.76 


-0.35 


1204-5U 


0.0318 


0.00 


3.14 


1303-07886 


0.0289 


0.00 


19.35 


1504-01 


0.0262 


2.81 


-2.39 


1504-11 


0.0210 


0.94 


-0.20 


1501-05882 


0.0194 


0.00 


11.54 


6024-0482 


0.0152 


1.01 


0.46 


1304-Sa 


0.0102 


0.00 


1.47 


13024-02*25 


0.0078 


0.00 


1.07 



1 


! 












05- 












■ m^ 00 ^ 


,M 


1 




■ 






a ^-*"^^ 
^^^^ a 


?» 










■ ^ 


^X"'b 


1 0-2- 

i 
0.1 


■ 

1 » 


a 


^^m ■ 






■ 

Waaftaiaa CNapoaa Am lamp** 





\" 


a 


% 






».0 07t2.0 0lM« R.071 



rnaqHn00gbyHydrooanPara.ua 
Flgura 1 • Weathered disposal m samples 




esq H nOOg by H »oiO| | en ParoiUa 

Ftjurelb. Fresh core samples 

Fgmsls-6 CwTr^r B r^^iouiiului(iijo»r4V»'l«*P<'^» l »o*r^««if>«^<>sp««^ 
ATM samples and rash cort simples on •» dntai sector) ■ TmiN Tonboon Wsanray 



1 








a 


a 




2 


■ 


■ 

a 


1 




■ 
a 




£ 

8 10. 

c 

X 










a 
^* s *»w^ a 




1 










- — 1 — — *■*» ■ 


Weatared Ohooaal 4iaa gampaa 
^^ y. 60487- 0.41 R-0.74 

WV » a aaa aa, 



Calculni(^n>or»lac<nil»alarrtac»*TnouaaiidTor»ll«larlal) 
Figure 2a. Weathered disposal Ansa samples. 




Fern* con Banska 

y.4»?< 0WIS> R.0.S1 



Calcium Carbonala Equtvatorrt (Tona / Thousand Tona Material) 
Figure 26 Fresh core samples 
Figures 2s-t> Cornpariton of potential adoVr, and calcium caitonste aqut»«l»nl on «e»lh»iad dispos«l 
ATM samples tnd lush con samples on t» cMOa sscoon • Tenrmssee Ttxttajfcee Wslerwsy 



208 



Results from the comparison of the 
acid-base accounting on the fresh core sam- 
ples are illustrated in figure 2b. There 
is high correlation between these methods 
(r^=0.81). However/ the same comparison 
on weathered disposal area soils shows a 
poor relationship between the two methods 
(r =0.42). Again/ this suggests that 
weathering products may be interfering with 
the peroxide procedure. 



Sobek/ A., W. Schuller/ J. Freeman and R. 
Smith. 1978. Field and Laboratory 
methods applicable to overburdens and 
minesoils. EPA-600/2-78-OS4. US-EPA, 
Cincinnati/ OH, pp. 204. 



CONCLUSIONS 



resu 

and 

weat 

rela 

that 

disp 

show 

erin 

ship 

and 

When 

soil 

Thes 

the 

acid 



A s 
Its 
tota 
here 
tion 

hav 
osal 
n th 
g f o 

exi 
tota 

pyr 
/ va 
e we 
pero 
ity 



tron 
obta 
1 su 
d ov 
ship 
e we 
are 
at i 
r lo 
sts 
1 su 
ite 
riou 
athe 
xide 
in s 



g re 
ined 
lfur 
erbu 

was 
athe 
as. 
n ma 
ng p 
betw 
lfur 
begi 
s we 
ring 

pro 
oil. 



lati 

usi 

by 
rden 

les 
red 

Pre 
teri 
erio 
een 

by 
ns t 
athe 

pro 
cedu 



onship 
ng per 
autoan 

core 
s pron 
follow 
vious 
al tha 
ds of 
the pe 
autoan 
o weat 
ring p 
ducts 
re for 



exist 
oxide 
alyzer 
sample 
ounced 
ing pi 
resear 
t has 
time, 
roxide 
alyzer 
her in 
roduct 
interf 

poten 



s betwe 

oxidati 

for un 

s. Thi 

on sam 

acement 

ch has 

been we 

no rela 

proced 

(WVU 1 

distur 

s evolv 

er with 

tial 



en 
on 

s 

pies 
in 

ath- 

tion- 

ure 

971). 

bed 

e. 



It was not within the scope of this 
study to examine the sources of error in 
the peroxide oxidation procedures; however/ 
future research could be directed toward 
identification of the interfering compounds 
and development of methods to reduce or 
eliminate these interferences. Based on 
the peroxide procedure be limited to fresh, 
unweathered samples. 

LITERATURE CITED 



West Virginia University/ 1971. Mine spoil 
potentials for water quality and con- 
trolled erosion. Water Pollution Con- 
trol Research Series. 14010 ESE 12/71 
US-EPA. Washington, DC, pp.206. 

Smith, R. M., W. E. Grube, Jr., T. Arkle, 

A. Sobek, 1974. Mine spoil potentials 
for soil and water quality. EPA-670- 
12-74-070. US-EPA. Cincinnati, OH, 
pp.303. 

Smith, R. M., A. Sobek, T. Arkle, Jr., J. 
C. Sencidiver, 1976. Extensive over- 
burden potentials for soil and water 
quality-EPA-600/2-76-184. US-EPA, 
Cincinnati, OH, pp. 311. 

Ammons, J. T., P. A. Shelton, and G. G. 

Davis, 1983. A detailed study of five 
overburden cores and six disposal 
areas along the divide section of the 
Tennessee-Tombigbee Waterway. U.S. 
Army Corps of Engineers, Nashville 
District, Nashville, TN. pp. 352. 



209 






THE INTERRELATION OF FACTORS INFLUENCING THE DISSOLUTION OF METALS 
IN COLUMNS OF MINE TAILINGS 1 

Richard D. Doepker' 1 



Abstract. — The dissolution of metal species from 
silver mine mill tailings containing quartz (70%) and 
manganif erous siderite was investigated using a 
column leach procedure. The effects of leachant 
ionic strength, pH, and buffering ability were exam- 
ined. The results indicated increased metal trans- 
port with increased ionic strength and decreased pH. 
In another test series, matched sets of columns were 
compared to determine the effect of leachate resi- 
dence time on metal concentration. In these tests, 
leachate metal ion concentrations more than doubled 
in columns that were allowed to evaporate (wet/dry 
cycle) about 30% of their pore leachate volume. The 
pH of the leachate eluted from the dry cycle (unsat- 
urated column) decreased while the conductance 
increased. Increases of five times were observed for 
lead and manganese, while increases of two orders of 
magnitude were seen for zinc and cadmium. The wet/ 
dry cycle effect was enhanced during subsequent 
cycles, but appeared not to be influenced by the 
addition of sodium lauryl sulfate, sodium benzoate, 
or phenol, which are known to inhibit microorganism 
involvement. Increased sensitivity to further oxi- 
dation of those columns that underwent the dry cycle 
became apparent through increased deviation of the 
ion yields from duplicate columns. The interrelation 
of the above factors with the composition of the 
tailings present adds complexity to determining dis- 
solution mechanisms. 



INTRODUCTION 

It is generally accepted that rain 
percolating through mine wastes can 
dissolve and transport metals through 



^Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and Recla- 
mation and the U.S. Department of the 
Interior (Bureau of Mines and Office of 
Surface Mining Reclamation and Enforce- 
ment). April 17-22, 1988, Pittsburgh, PA. 

2 Richard D. Doepker is a research 
chemist, Spokane Research Center, U.S. 
Bureau of Mines, Spokane, WA. 



the wastes and into the surrounding area. 
Depending on the toxicity of the metals 
in question and their relative concentra- 
tions, such a situation could lead to 
contamination of surface and ground 
waters, rendering them toxic to plants, 
fish, and animals. The question is, 
therefore, how can it be determined 
whether an individual site contains haz- 
ardous material? 

The methods presently used to pre- 
dict acid mine drainage (Ferguson 1987) , 
although not completely successful, do 
offer one factor that may well influence 
the concentrations of leached metal 
species, and that is acidity. Studies by 
Calspan (Brainbridge 1980) and A. D. 



210 



Table 1. — Assay of silver mine tailings, pet. 



Si 


Fe 


Al 


C 


Mn 


Pb 


Ca 


Mg 


S 


Zn 


K 


Cu 


As 


35.4 


6.6 


2.4 


1.8 


0.65 


0.43 


0.25 


0.25 


0.16 


0.11 


0.08 


0.04 


0.02 



Little, Inc. (Kuryk 1985) , have produced 
additional information and protocols for 
column leach studies that may permit 
controlled laboratory experiments to 
intersect, divide, and isolate those 
kinetic and thermodynamic processes that 
influence the environmental mobilization 
of metal species through and out of mine 
mill tailings. 

The laboratory program for this 
project involved a series of tests on 
mill tailings from a silver mine. We 
examined the effects of leachant ionic 
strength, leachant acidity, the presence 
of acetate ions, and the influence of a 
"wet and dry cycle" on the dissolution of 
metals. Analyses were performed on the 
leachates for aluminum (Al) , calcium 
(Ca) , cadmium (Cd) , copper (Cu) , iron 
(Fe) , lead (Pb) , manganese (Mn) , mag- 
nesium (Mg) , nickel (Ni) , potassium (K) , 
sodium (Na) , and zinc (Zn) . Ionic con- 
ductivity and pH were also determined 
for each leachate. The purpose of the 
study was to isolate certain parameters 
and to determine their kinetic effects on 
the overall transport of metal species 
within the tailings sample. 

MATERIALS AND METHODS 

Waste 

The silver mine tailings used in 
this investigation contained partially 
liberated quartz (70%) and manganif erous 
siderite (FeCC^-Mn) as principal gangue 
minerals. The quartz and chert ranged 
from 10 to 15 micrometers and the sid- 
erite from 10 to 100 micrometers. The 
principal metallic minerals were pyrite 
(FeS 2 ) , galena (PbS) , sphalerite (ZnS) , 
minor chalcopyrite (CuFeS2) , minor 
tetrahedrite [ (Cu,Fe,Zn,Ag) i2 sb 4 s i3l ' 
and minor muscovite [KAl-jSijO^g (OH) 9] . 
Particles of these minerals ranged from 
1 to 20 micrometers in diameter. A 
trace of digenite (CU2S) and possibly 
arsenopyrite (FeAsS) were detected by 
X-ray diffraction. The pyrite was 
generally not attached to other minerals 
and had shiny unaltered surfaces even 
though the tailings were weathered. A 
few pyrite grains had inclusions of 
galena and chalcopyrite. About one-half 
of the galena was completely included 
within siderite or, rarely, quartz. Most 
of the remaining galena was attached to 
siderite grains but had exposed surfaces; 
only a minor portion of the galena was 
free. Sphalerite occurred as large 
unattached grains or smaller grains 
attached to siderite. Chalcopyrite and 
tetrahedrite most commonly occurred 



together as small inclusions in siderite 
or galena. A partial listing of the 
results of the destructive assay of the 
tailings is given in table 1. 

The tailings samples used for all 
experiments were oven dried at 160° F in 
half-filled 55-gal drums. The material 
was crushed to its original size range, 
mixed well, and stored in large plastic 
drums. 

Chemicals 

All chemicals in this study were 
commercially available, analytical-grade 
reagents (A.R. grade) used without fur- 
ther purification. The deionized water 
was produced in the laboratory through 
distillation (Barnstead glass still) then 
deionized with a Barnstead NANOpure II 
Demineralizer 3 (18.8 megohm-cm). Leach- 
ant solutions were prepared by standard 
analytical techniques using only A.R. 
grade chemicals and prepared deionized 
water. They were then stored in carboys 
(Nalgene) until used. 

Column Test Equipment and Methods 

Leach columns were constructed 
from 2- or 4-ft lengths of 3-in-inside- 
diameter (ID) PVC pipe equipped with 
cemented couplings and bushings in which 
perforated Nalgene plates had been 
installed. A 9-cm G6 (Fisher Scientific) 
borosilicate glass fiber filter was 
placed on the perforated plate before 
installation of the bushing. Twenty of 
the 2-ft columns were mounted in a rack 
and arranged in groups of four, permit- 
ting a test sequence for five parameters. 
Combinations of other 2- and 4-ft columns 
were used for various specific experi- 
ments. A series of 10 similarly con- 
structed 1-1/2-in-ID PVC columns were 
used to examine the effects of sample 
depth and column surface-to-volume 
ratios. These columns were arranged in 
series of two, again for a five-parameter 
study. Leachants were introduced to the 
columns drop by drop from a 1-liter 
polypropylene storage bottle through 
tygon tubing fitted with a screw clamp. 
Volumes of leachant used depended on an 
experimentally determined pore volume. 
Estimates of pore volume were made by 
weighing out 100 g of tailings, satu- 
rating the sample with deionized water, 
and placing it in a Buchner funnel. The 



^Reference to specific products or 
manufacturers does not imply endorsement 
by the Bureau of Mines. 



211 



sample was 
overnight , 
difference 
weight of w 
the sample, 
test series 
appropriate 
leachant to 
additional 
between the 
gave a corr 
experiments 
equal to or 
pore volume 



covered, allowed to drip-dry 
and then reweighed. The 
in weight represented the 
ater stored in the pores of 
The columns to be used in a 
were then filled with the 
weight of tailings and enough 
saturate the column plus an 
150 cm . The difference 

leachant and leachate volumes 
ected pore volume. In all 
, the leachant introduced was 
a specific fraction of the 



Slurry Test Methods 



Where slu 
weighed sample 
into a 500 cm 3 
bottle (Nalgen 
duced in prede 
ratios. The b 
variable-speed 
4 to 16 hours, 
filtered throu 
paper. The pH 
leachate was d 
leachate was a 
acid for analy 
techniques or 
analysis with 



rry procedures were used, a 
of the tailings was placed 
polypropylene rectangular 
e) . Leachant was intro- 
termined solid-to-liquid 
ottles were placed on a 
, reciprocating shaker for 

The samples were then 
gh no. 42 Whatman filter 
and conductivity of the 
etermined, after which the 
cidified to 0.01M. nitric 
sis by atomic absorption 
to about 2% nitric acid for 
ICP. 

Equipment 



Most of the metal analyses reported 
in this investigation were carried out 
with the aid of an IL 551 atomic absorp- 
tion spectrometer. With the acquisition 
of a Perkin-Elmer Plasma II, ICP became 
the method of choice for metal analyses. 

DISCUSSION OF RESULTS 

Ionic Strength Effect 



Tables 2-A and 
concentrations for 1 
ionic strengths and 
mental method utiliz 
of tailings samples 
per each column; fou 
different leachants 
0.10M KN0 3 ; deionize 
of Western synthetic 
1980) . The pore vol 
was determined to be 
pH was adjusted to a 
(table 2-A) or 2.0 ( 
nitric acid. 



2-B report leachate 
eachants of various 
pK's. The experi- 
ed 3-in columns; 2 kg 
(10.5-11 in, depth) 
r replicates; five 
of 0.001M., 0.01M., and 
d water; and a sample 

rain (Bainbridge 
ume for this series 

500 cm 3 . Leachant 
pproximately 5.5 
table 2-B) with 



It should be noted that 
demonstrated a rapid decreas 
tration from that shown in t 
saturation experiment (test 
two to three pore volumes of 
were added. After this init 
leachate concentrations reac 
constant value for all speci 
nearly all cases, the concen 
metal contaminants was great 
the O.IOM.KNO3 leachant. In 
the magnitude of this increa 
tration was found to be due 



all species 
e in concen- 
he initial 
84) when from 

leachant 
ial decrease, 
hed a near- 
es. In 
tration of 
est for 

general, 
sed concen- 
to changes in 



solubility as a result of changes in 
activity coefficients. Unfortunately, 
the increased leachate pH of nearly 
0.4 was above similar expectations. 
Furthermore, it was expected that this 
increased leachate pH would decrease 
metal dissolution, resulting in a com- 
bined effect of little to no apparent 
change in ion concentration. 

With the adjustment of the leachant 
pH to a value of 2.0 using nitric acid, 
metal dissolution increased while the 
solution maintained the previous ionic 
strength effect and the leachate pH 
decreased to a value between 4.1 and 4.5. 
This sample of silver mill tailings 
appeared to have a strong buffer region 
between a pH of 4.5 to 3.5. The pH of a 
20:1 liquid:solid slurry formed with 
deionized water and the silver mill tail- 
ings when varying amounts of sulfuric 
acid were added reproduced this buffer 
region (fig. 1) . Table 2-B shows that 
leachate concentrations for most metal 
species (tests 012 through 016) declined 
after the initial influence of the acidic 
leachant, and that repeated "washing" 
with deionized water returned the system 
to the pre-acid leaching condition. 
Nickel concentrations in the leachates 
were found to be below detection limits 
(0.06 ppm) until the pH 2 leachant was 
used. A value of 1.06 ppm was observed 
for nickel concentrations of 0.10M, KNO3, 
0.86 ppm for 0.01M. KNO3 , and 0.82 for 
0.001M KNO3 (test 013). Initial cadmium 
leachate concentrations (test 84) were 
0.39, 0.29, and 0.27 ppm for O.lOtL 
0.01M, and 0.001M KNOo, respectively. 
These values decreased to below detection 
limits (0.01) for all leachant systems 
except the 0.10M. KNOo leachant (at 0.021) 
after the addition of two pore volumes of 
leachant (test 90) . Cadmium concentra- 
tions remained below detection limits for 
all leachings until the acid leachant 
effect was observed, whereupon concentra- 
tions of 0.04 to 0.02 ppm were noted. 



8 




1 1 1 1 1 1 1 


I 7 


- \g 


_ 


a. 






y, 


\° 




2 






< fi 




- 


I 






O 






< 






ID 






-J 5 






4 




j Q 




■ 






-2 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 

ADDED ACID. 10- 2 m mol(H+)/g sample 



Figure 1. — Leachate pH as a function of 
added acid. 



212 



To further test the ionic strength 
effect, newly prepared columns containing 
1.5 kg of the silver mill tailings were 
first leached with deionized water and 
then subjected to one of five leachants 
(0.10MKC1, 0.033M K 2 S0 4 , O.lO&KNOo, 
0.033M. Na 2 S0 4 , or . I0CL LiN0 3 ) of iden- 



tical ionic strengths. The results of 
the tests are summarized in table 3 
and show that, in general, the ionic 
strength effect previously observed was 
independent of the salt used to control 
the ionic strength. 



TABLE 2-A IONIC STRENGTH EFFECT 

Ag-MINE TAILINGS 3-inch (i.d.) column; 2.00 kg in each; height "10.5-11 inches. 

AVERAGE (4 columns) 



# 




84 


86 


88 


90 


92 


94 


96 


98 


100 


104 


110 


114 


Resident time(hr) 


start 


72 


96 


48 


48 


48 


96 


96 


48 


288 


48 


288 


Leachant volume 




250 


250 


250 


250 


250 


250 


250 


250 


500 


500 


500 


500 


__l FATHANT-. 




























PH 




























DW, pH = 5.40 




6.90 


7.07 


7.23 


7.41 


7.36 


7.34 


7.44 


7.61 


7.44 


7.47 


7.52 


7.66 


0.001M,KN03,pH=! 


i.46 


6.98 


7.10 


7.23 


7.43 


7.35 


7.48 


7.56 


7.69 


7.55 


7.44 


7.56 


7.72 


0.01M,KN03,pH=5. 


43 


6.94 


7.12 


7.23 


7.26 


7.29 


7.50 


7.60 


7.75 


7.71 


7.72 


7.76 


7.86 


0.10M,KN03,pH=5. 


61 


6.92 


7.15 


7.21 


7.31 


7.51 


7.96 


8.06 


8.03 


8.02 


7.94 


7.99 


8.00 


RAIN,(SYN),pH=4. 


50 


6.96 


7.12 


7.25 


7.38 


7.38 


7.54 


7.60 


7.67 


7.48 


7.50 


7.60 


7.66 


Calcium (Ca), mg/L 




























DW 




590 


627 


598 


95.9 


29.7 


22.6 


25.3 


22.4 


17.0 


22.6 


16.9 


19.5 


0.001M.KN03 




630 


668 


606 


97.5 


28.9 


23.6 


26.3 


22.7 


16.4 


22.3 


18.2 


21.6 


0.01M.KN03 




680 


774 


623 


91.3 


33.5 


26.5 


28.7 


24.2 


17.9 


23.9 


21.5 


22.8 


0.10M.KN03 




982 


958 


406 


48.4 


39.5 


32.8 


34.2 


31.5 


20.2 


31.1 


21.0 


28.4 


RAIN.(SYN) 




643 


678 


574 


76.4 


27.4 


21.4 


24.3 


20.1 


14.2 


22.6 


16.6 


19.8 


Conductance, mmho. 




























DW, 0.0036 mmho 


3.68 


2.51 


2.22 


1.952 


0.197 


0.162 


0.178 


0.150 


0.126 


0.166 


0.125 


0.139 


0.001M.KN03, 0.156 


3.70 


2.56 


2.31 


0.648 


0.302 


0.289 


0.311 


0.288 


0.265 


0.312 


0.257 


0.274 


0.01M.KN03, 1.343 


4.53 


3.74 


3.43 


1.72 


1.46 


1.48 


1.48 


1.46 


1.45 


1.50 


1.45 


1.48 


0.10M.KN03, 11. 


54 


13.2 


13.2 


11.7 


10.7 


11.0 


11.5 


11.7 


11.6 


11.6 


11.7 


11.8 


11.8 


RAIN.(SYN), 0.0144 


3.59 


2.59 


2.16 


0.434 


0.186 


0.156 


0.174 


0.146 


0.123 


0.164 


0.122 


0.138 


Iron (Fe), mg/L 




























DW 




0.08 


0.05 


0.05 


* 


0.15 


* 


* 


* 


* 


* 


* 


* 


0.001M.KN03 




0.07 


0.04 


0.05 


* 


0.92 


0.12 


* 


* 


0.11 


* 


* 


* 


0.01M.KN03 




0.08 


0.04 


0.05 


* 


0.13 


0.30 


* 


* 


0.80 


* 


* 


0.06 


0.10M.KN03 




0.11 


0.08 


0.08 


* 


0.05 


* 


0.04 


* 


* 


+ 


+ 


0.04 


RAIN.(SYN) 




0.08 


0.07 


0.04 


0.04 


0.04 


* 


* 


* 


+ 


0.04 


* 


* 


* below dl, dl=0 

Magnesium (Mg), mg/L 


04 




















































DW 




145 


12.7 


2.72 


2.86 


2.44 


2.28 


2.90 


2.23 


1.99 


3.21 


1.97 


2.50 


0.001M.KN03 




150 


11.8 


3.27 


2.44 


2.55 


2.45 


2.85 


2.09 


2.02 


2.80 


1.90 


2.38 


0.01M.KN03 




148 


13.7 


6.62 


4.32 


3.49 


2.88 


3.36 


2.37 


2.39 


3.07 


2.44 


2.64 


0.10M.KN03 




144 


16.8 


13.13 


5.01 


3.85 


3.30 


4.20 


3.15 


2.88 


4.75 


2.68 


2.80 


RAIN.(SYN) 




138 


11.5 


2.65 


2.10 


2.22 


1.94 


2.34 


1.70 


1.82 


3.04 


1.88 


2.43 


Lead (Pb), mg/L 




























DW 




3.38 


2.26 


1.50 


0.49 


0.34 


0.31 


0.24 


0.18 


0.16 


0.24 


0.17 


0.17 


0.001M.KN03 




3.34 


2.42 


1.64 


0.49 


0.39 


0.34 


0.27 


0.24 


0.13 


0.26 


0.17 


0.25 


0.01M.KN03 




3.60 


2.55 


1.67 


0.63 


0.51 


0.42 


0.33 


0.29 


0.17 


0.25 


0.14 


0.24 


0.10M.KN03 




6.88 


4.07 


1.90 


2.90 


0.67 


0.52 


0.41 


0.26 


0.20 


0.28 


0.17 


0.21 


RAIN.(SYN) 




4.21 


2.67 


1.87 


0.48 


0.32 


0.27 


0.27 


0.25 


0.19 


0.25 


0.19 


0.19 


* below dl, dl=0 

Zinc (Zn), mg/L 


10 




















































DW 




21.3 


7.35 


3.48 


0.638 


0.218 


0.164 


0.160 


0.138 


0.109 


0.185 


0.150 


0.267 


0.001M.KN03 




20.3 


9.51 


3.59 


0.636 


0.254 


0.201 


0.197 


0.157 


0.114 


0.284 


0.193 


0.325 


0.01M.KN03 




22.9 


9.56 


4.89 


1.31 


0.679 


0.440 


0.294 


0.226 


0.174 


0.242 


0.195 


0.328 


0.10M.KN03 




33.1 


15.4 


8.33 


2.58 


1.00 


0.506 


0.292 


0.240 


0.193 


0.277 


0.235 


0.440 


RAIN.(SYN) 




19.9 


7.55 


3.60 


0.519 


0.210 


0.168 


0.151 


0.133 


0.114 


0.231 


0.152 


0.259 



DW - deionized water 



213 



TABLE 2-B IONIC STRENGTH EFFECT 
Ag-MINE TAILINGS 3-inch (i.d.) co 
AVERAGE (4 columns) 



umn; 2.00 kg in each; height "10.5-11 inches. 





008 




010 


012 


013 


014 




016 


018 


019 


021 


022 


Resident time(h 


48 




168 


168 


216 


336 




120 


216 


168 


168 


456 


Leachant volume 


500 




500 


500 


500 


500 




500 


500 


500 


500 


500 


LEACHANT-- 




---LEACHANT- 










LEACHANT- 












pH 




pH = 2.00 






















DW 


7.79 


DW 


7.65 


4.65 


4.17 


4.35 


DW 


4.77 


6.56 


7.27 


7.53 


7.66 


0.001M.KN03 


7.80 


0.001M.KN03 


7.71 


4.56 


4.13 


4.40 


DW 


4.81 


6.91 


7.48 


7.65 


7.69 


0.01M.KN03 


7.99 


0.01M.KN03 


7.84 


4.71 


4.21 


4.39 


DW 


4.92 


7.39 


7.70 


7.75 


7.72 


0.10M.KN03 


8.21 


0.10M.KN03 


8.06 


4.79 


4.25 


4.43 


DW 


5.04 


7.92 


7.90 


7.85 


7.82 


RAIN.(SYN) 


7.78 


RAIN.(SYN) 
■ — LEACHANT— 


7.64 


4.72 


4.15 


4.30 


DW 
LEACHANT- 


4.87 


6.73 


7.58 


7.71 


7.65 


Ca, mg/L 




pH = 2.00 






















DW 


16.6 


DW 


55.0 


137 


72.5 


70.2 


DW 


36.4 


10.2 


13.7 


14.2 


24.7 


0.001M.KN03 


17.7 


0.001M.KN03 


47.5 


137 


74.0 


55.2 


DW 


34.4 


9.74 


12.8 


14.4 


24.6 


0.01M.KN03 


19.2 


0.01M.KN03 


33.5 


118 


68.1 


71.0 


DW 


36.8 


11.2 


13.6 


15.4 


25.9 


0.10M.KN03 


20.4 


0.10M.KN03 


28.2 


103 


76.0 


64.6 


DW 


39.7 


4.06 


10.5 


14.2 


32.6 


RAIN.(SYN) 


16.7 


RAIN.(SYN) 
— LEACHANT— 


56.3 


118 


65.9 


56.3 


DW 
LEACHANT- 


31.7 


7.93 


11.7 


13.8 


23.7 


Conductance, mmhc 




pH = 2.00 






















DW 


0.123 


DW 


0.421 


2.11 


2.18 


nd 


DW 


1.97 


0.125 


0.105 


0.113 


0.165 


0.001M.KN03 


0.261 


0.001M.KN03 


0.471 


2.21 


2.27 


nd 


DW 


2.00 


0.147 


0.116 


0.119 


0.169 


0.01M.KN03 


1.47 


0.01M.KN03 


1.79 


3.38 


3.40 


nd 


DW 


2.98 


0.282 


0.150 


0.139 


0.183 


0.10M.KN03 


11.9 


0.10M.KN03 


12.3 


13.5 


13.5 


nd 


DW 


11.9 


0.714 


0.234 


0.162 


0.262 


RAIN.(SYN) 


0.121 


RAIN.(SYN) 
— LEACHANT— 


0.433 


2.10 


2.22 


nd 


DW 
LEACHANT- 


1.96 


0.129 


0.108 


0.113 


0.162 


Fe, mg/L 




pH = 2.00 






















DW 


* 


DW 


* 


219 


336 


337 


DW 


321 


16.6 


9.14 


4.43 


0.14 


0.001M.KN03 


* 


0.001M.KN03 


* 


181 


321 


314 


DW 


330 


12.9 


5.39 


3.93 


0.18 


0.01M.KN03 


* 


0.01M.KN03 


* 


195 


325 


290 


DW 


307 


6.69 


3.83 


2.31 


0.19 


0.10M.KN03 


* 


0.10M.KN03 


0.14 


173 


323 


324 


DW 


310 


0.88 


1.29 


1.42 


* 


RAIN.(SYN) 


* 


RAIN.(SYN) 


0.05 


203 


354 


354 


DW 


318 


13.0 


7.17 


3.79 


0.24 


* below dl, dl=0 


.04 


--LEACHANT— 










LEACHANT- 












Mg, mg/L 




pH = 2.00 






















DW 


2.13 


DW 


6.06 


44.3 


22.4 


24.8 


DW 


22.3 


2.52 


1.71 


1.66 


2.33 


0.001M.KN03 


2.07 


0.001M.KN03 


5.05 


38.1 


21.7 


24.8 


DW 


19.6 


2.26 


1.63 


1.60 


2.48 


0.01M.KN03 


2.40 


0.01M.KN03 


3.80 


47.9 


22.4 


26.4 


DW 


18.8 


2.13 


1.74 


1.77 


2.57 


0.10M.KN03 


2.61 


0.10M.KN03 


3.52 


46.5 


23.8 


26.7 


DW 


19.0 


0.77 


0.95 


1.17 


3.06 


RAIN.(SYN) 


1.98 


RAIN.(SYN) 
— LEACHANT— - 


7.54 


58.5 


22.3 


23.6 


DW 
LEACHANT-- 


19.3 


2.05 


1.62 


1.50 


2.35 


Pb. mg/L 




pH = 2.00 






















DW 


0.13 


DW 


0.27 


13.2 


3.59 


2.67 


DW 


0.41 


0.20 


* 


* 


* 


0.001M.KN03 


0.12 


0.001M.KN03 


0.40 


8.63 


4.16 


2.37 


DW 


0.21 


* 


* 


+ 


* 


0.01M.KN03 


0.11 


0.01M.KN03 


0.23 


10.4 


6.50 


2.77 


DW 


1.36 


* 


* 


* 


+ 


0.10M.KN03 


0.16 


0.10M.KN03 


0.18 


8.65 


1.65 


0.33 


DW 


0.17 


* 


* 


0.17 


* 


RAIN.(SYN) 


0.15 


RAIN.(SYN) 


0.21 


10.9 


1.58 


2.43 


DW 


0.20 


0.11 


* 


* 


* 


* below dl, dl=. 


910 


— LEACHANT— - 











uEACHANT- 












Zn, mg/L 




pH = 2.00 






















DW 


0.211 


DW 


0.886 


15.0 


19.2 


32.7 


DW 


31.4 


3.3 


1.3 


0.9 


0.9 


0.001M.KN03 


0.209 


0.001M.KN03 


0.615 


16.8 


18.7 


30.7 


DW 


31.4 


2.9 


1.1 


0.9 


0.9 


0.01M.KN03 


0.249 


0.01M.KN03 


0.416 


22.4 


20.3 


39.9 


DW 


34.8 


2.0 


1.1 


0.9 


0.8 


0.10M.KN03 


0.290 


0.10M.KN03 


0.334 


28.8 


26.0 


45.6 


DW 


38.8 


0.5 


0.7 


0.8 


0.8 


RAIN.(SYN) 


0.199 


RAIN.(SYN) 


0.787 


17.7 


18.5 


35.8 


DW 


30.9 


2.8 


1.4 


0.9 


0.8 



DW - deionized water 



Residence Time, Wet/Dry Cycle Effect 

Tables 2 and 3 indicate that the 
leachate ion concentrations appeared to 
be influenced by the residence time, that 
is, the time while the leachate remained 
in the column. A close inspection of 



these tables reveals that the concentra- 
tions of most species increased with 
increasing residence time, with the pos- 
sible exception of lead (table 3) (test 
055 compared to 063) . In order to exam- 
ine this effect further, the columns 
used in the "salt" study (table 3) were 



214 



TABLE 3 SALT EFFECT 






















Ag-MINE TAILINGS VIII 3- 


inch (i. 


d.) column; 1.50 kg in 


each; leachant 


volume. 


350cc. 










AVERAGE (4 columns 
# 


) 

032 


035 




037 


042 


045 


049 


055 


059 


063 


068 


Resident time(hr) 


start 


144 




168 


168 


168 


168 


366 


192 


72 


168 


— LEACHANT- 

PH 

DW 






LEACHANT 
0.10M,KCl,pH=5.49 


















6.47 


6.49 


7.24 


7.30 


7.61 


7.73 


7.87 


7.93 


7.96 


7.97 


DM 


6.53 


6.25 


0.033M,K2S04,pH=5.59 


7.28 


7.43 


7.82 


7.95 


8.15 


8.08 


8.15 


8.10 


DW 


6.28 


6.96 


0.10M,KN03,pH=5.39 


7.27 


7.39 


7.72 


7.86 


7.99 


7.96 


8.01 


8.02 


DW 


6.57 


7.00 


0.033M,Na2S04,pH=5.63 


7.57 


7.46 


7.79 


7.91 


8.06 


8.07 


n.d. 


8.14 


DW 


6.33 


7.13 


0.10M,LiN03,pH=5.58 


7.47 


7.55 


7.74 


7.85 


7.95 


7.95 


n.d. 


7.98 


— n.d. not determined 
CALCIUM (Ca), mg/L 

DW 655 






















277 


0.10M, KC1 


149 


54.5 


35.8 


27.8 


33.1 


26.2 


17.0 


28.3 


DW 


672 


239 


0.033M, K2S04 


98.8 


64.4 


35.7 


26.9 


35.9 


28.5 


_20.8 


34.7 


DW 


656 


239 


0.10M, KN03 


123 


60.2 


33.6 


25.3 


27.5 


25.5 


17.7 


31.1 


DW 


635 


231 


0.033M, Na2S04 


80.4 


65.5 


30.6 


27.1 


33.1 


31.8 


26.1 


25.3 


DW 


687 


178 


0.10M, LiN03 


74.4 


59.2 


32.2 


24.5 


28.8 


29.8 


26.4 


29.5 


CONDUCTANCE, mmho 
























DW 


2.72 


1.04 


0.10M,KCl,12.58mmho 


2.16 


12.27 


12.92 


13.17 


13.26 


12.76 


13.13 


13.04 


DW 


2.73 


0.99 


0.033M,K2S04,7.53mmho 


1.18 


7.35 


7.84 


8.39 


8.09 


7.76 


7.91 


8.00 


DW 


2.74 


1.04 


0.10M,KN03,11.80mmho 


2.23 


11.57 


12.22 


12.46 


12.63 


12.05 


12.19 


12.34 


DW 


2.69 


0.98 


0.033M,Na2S04,6.20mmho 


1.24 


6.25 


6.55 


7.13 


6.81 


6.49 


n.d. 


6.57 


DW 


2.70 


0.82 


0.10M,LiN03,8.95mmho 


1.79 


8.85 


9.09 


9.40 


9.33 


8.87 


n.d. 


8.95 


— n.d. not determined 
IRON (Fe), mg/L 

DW 0.13 






















0.09 


0.10M, KC1 


0.14 


0.17 


0.09 


0.06 


0.08 


0.05 


0.10 


0.16 


DW 


0.12 


0.07 


0.033M, K2S04 


0.22 


0.08 


<.04 


0.05 


0.04 


0.05 


0.14 


0.11 


DW 


0.24 


0.11 


0.10M, KN03 


0.37 


0.06 


0.05 


0.04 


0.05 


<.04 


0.27 


0.09 


DW 


0.08 


0.19 


0.033M, Na2S04 


0.08 


0.18 


0.17 


<.04 


0.07 


<.04 


0.16 


0.14 


DW 


0.12 


0.08 


0.10M, LiN03 


0.11 


0.25 


0.15 


0.08 


0.07 


0.08 


0.14 


0.22 


MAGNESIUM (Mg), mg/L 

DW 43.3 


5.07 


0.10M, KC1 


12.2 


5.77 


3.02 


3.00 


4.41 


3.28 


2.57 


3.29 


DW 


41.0 


4.98 


0.033m, K2S04 


11.3 


6.19 


2.64 


2.80 


4.28 


3.77 


2.78 


3.58 


DW 


43.7 


4.81 


0.10M, KN03 


12.3 


5.63 


2.96 


2.86 


3.91 


3.28 


2.72 


3.21 


DW 


41.6 


5.01 


0.033M, Na2S04 


7.64 


6.13 


2.49 


2.37 


3.33 


3.38 


3.14 


3.50 


DW 


40.3 


4.26 


0.10M, LiN03 


7.44 


5.67 


2.46 


2.60 


3.09 


2.88 


3.15 


5.16 


LEAD (Pb), mg/L 
DW 


2.86 


1.14 


0.10M, KC1 


0.64 


0.76 


0.37 


0.28 


0.30 


0.20 


2.45 


0.34 


DW 


3.12 


0.66 


0.033M, K2S04 


0.63 


0.69 


0.34 


0.23 


0.29 


0.24 


1.27 


1.30 


DW 


2.93 


0.78 


0.10M, KN03 


0.61 


0.43 


0.43 


0.34 


0.24 


0.14 


10.7 


0.48 


DW 


2.94 


1.20 


0.033M, Na2S04 


0.42 


0.44 


0.20 


0.29 


0.25 


0.26 


0.23 


1.08 


DW 


3.38 


0.82 


0.10M, LiN03 


0.39 


0.34 


0.33 


0.26 


0.26 


0.19 


3.68 


n.d. 


_-_ n r\ nnt HptprminoH 






















ZINC (Zn), mg/L 
DW 


12.8 


2.50 


0.10M, KC1 


1.55 


1.52 


0.61 


0.44 


0.84 


0.76 


0.58 


0.55 


DW 


13.4 


1.75 


0.033M, K2S04 


1.07 


1.12 


0.58 


0.42 


0.69 


0.64 


0.45 


0.42 


DW 


13.3 


1.89 


0.10M, KN03 


1.57 


1.20 


0.41 


0.37 


0.52 


0.44 


0.32 


0.42 


DW 


13.8 


2.57 


0.033M, Na2S04 


0.76 


0.77 


0.63 


0.52 


0.75 


0.42 


0.33 


0.59 


DW 


13.4 


2.29 


0.10M, LiN03 


1.12 


0.92 


0.51 


0.41 


0.45 


0.40 


0.27 


0.40 



DW - deionized water 



divided into pairs, two of which were 
maintained at near-saturation conditions 
while the other was allowed to dry 
partially. The averages of the two 
saturated (even-numbered) and unsaturated 
(odd-numbered) columns before and after 
the wet/dry cycle are presented in table 
4. It becomes apparent that those 



columns allowed to evaporate approxi- 
mately 30% to 40% of their pore liquid 
produced an environment that resulted in 
increased metal dissolution and decreased 
leachate pH. After stabilization (addi- 
tion of three pore volumes of "salt" 
leachant) , the process was repeated. The 
results were similar. Presumably, this 



215 



was because of atmospheric oxidation of 
the sulfide minerals present in the 
unsaturated zone and the subsequent for- 
mation of acid (Nordstrom et al. 1979, 
Nordstrom 1982) . In order to examine a 
possible relation between microorganism 
activity and oxidation, sodium lauryl 
sulfate, sodium benzoate, or phenol 
(Watzlaf 1986) were added to three sets 



of columns, while 0.01M. KNO3 and 0.01M. 
KN0 2 leachants were used on the remain- 
ing two sets. Although the presence or 
absence of microorganisms was not experi- 
mentally established, it was assumed that 
the treatment should have eliminated them 
from contributing to the next wet/dry 
cycle experiment. It may be noted that 
this test (test 104) produced effects 



Ag-MINE TAILINGS 


3-inch 


(i.d.) 


column; 


1.50 kg 


in each; 


leachant volume, 


350 cc 










AVERAGE (2 columns) 
























# 


059 


068 




073 


075 


080 


083 


088 




095 


104 


108 


Resident time(hr) 


192 


168 




504 


168 


168 


504 


168 




336 


576 


120 










SAT/UNSAT 




SAT/UNSAT 








SAT/UNSAT 


AVG. LEACHANT 






AVG. 








LEACHANT 


LEACHANT 






LEACHANT 


nH KT1 
















DW(ALL) 


SLS 






DW(ALL) 














28,4 (0.10M) 


7.92 


7.98 


SAT. 


7.91 


8.00 


8.07 


8.03 


8.05 


550mg/L 


7.88 


7.89 


7.74 


18,3 pH=5.49 


7.94 


7.97 


UNSAT. 


7.25 


7.81 


7.96 


7.31 


7.72 


pH=5.40 


7.78 


7.32 


7.70 


K oc ;nd 


















NaBz 
578mg/L 








68,8 (0.033M) 


8.09 


8.11 


SAT. 


8.09 


8.15 


8.17 


8.12 


8.21 


8.00 


8.10 


7.91 


5&7 pH=5.59 


8.06 


8.09 


UNSAT. 


7.80 


8.01 


8.07 


7.55 


7.96 


pH=6.13 


7.92 


7.18 


7.74 


KN03 


















KN03 
0.01M 








108.12 (0.10M) 


7.90 


8.03 


SAT. 


8.09 


7.89 


8.07 


7.98 


8.02 


7.82 


7.97 


8.01 


98.11 pH=5.39 


8.02 


8.01 


UNSAT. 


7.51 


7.85 


8.00 


6.92 


7.82 


pH=5.40 


7.71 


5.00 


6.99 


Na2S04 - 
148.16 (0.033M) 


















PHENOL 
578mg/L 








8.11 


8.15 


SAT. 


8.06 


8.13 


8.18 


8.16 


8.30 


7.82 


8.00 


7.97 


138.15 pH=5.63 


8.04 


8.13 


UNSAT. 


6.95 


7.87 


8.19 


7.38 


8.03 


pH=5.10 


8.02 


7.97 


8.04 


1 iNfn 


















NaN02 
0.01M 








188.20 (0.10M) 


7.94 


8.00 


SAT. 


7.98 


7.89 


8.07 


7.93 


8.06 


7.95 


7.96 


7.97 


178,19 pH=5.58 


7.95 


7.97 


UNSAT. 


7.55 


7.82 


7.98 


7.18 


7.94 


pH=6.06 


7.74 


5.66 


7.20 


CALCIUM (Ca), mg/L 


























28.4 KCl 


27.80 


29.30 


SAT. 


20.30 


26.95 


21.15 


23.80 


18.10 


SLS 


1.31 


1.53 


0.89 


18.3 (0.10M) 


24.70 


27.40 


UNSAT. 
SAT. 


53.75 


28.30 


21.85 


112.25 


27.20 


550mg/L 
NaBz 


6.55 


47.9 


10.9 


68.8 K2S04 


28.50 


38.00 


23.05 


31.35 


24.00 


29.20 


18.45 


2.53 


4.18 


4.98 


58.7 (0.033M) 


29.60 


31.50 


UNSAT. 
SAT. 


47.60 


27.90 


23.05 


68.30 


18.20 


578mg/L 
KN03 


2.75 


59.5 


12.9 


108.12 KN03 


25.90 


35.90 


17.95 


22.60 


16.35 


24.30 


14.90 


13.0 


17.2 


10.5 


98,11 (0.10M) 


25.20 


26.30 


UNSAT. 
SAT. 


50.30 


23.80 


18.40 


81.75 


15.15 


0.01M 
PHENOL 


18.8 


101 


18.6 


148.16 Na2S04 


33.30 


24.70 


21.30 


32.05 


22.45 


26.55 


23.80 


12.0 


14.6 


13.6 


138,15 (0.033M) 


30.30 


26.00 


UNSAT. 
SAT. 


96.75 


37.85 


31.85 


232 


52.70 


578mg/L 
KN02 


17.1 


81.3 


30.7 


188,20 LiN03 


29.30 


31.60 


21.30 


37.60 


24.60 


31.20 


22.95 


11.8 


15.7 


12.3 


178,19 (0.10M) 


30.30 


27.40 


UNSAT. 


56.10 


21.40 


27.50 


90.75 


18.00 


0.01M 


20.7 


82.9 


15.4 


MAGNESIUM (Mg), mg/L 
























28,4 KCl 


3.43 


3.30 


SAT. 


3.16 


3.84 


2.71 


1.30 


2.26 


SLS 


0.15 


0.27 


0.11 


18,3 (0.10M) 


3.14 


3.27 


UNSAT. 
SAT. 


10.2 


4.02 


4.12 


11.9 


2.90 


550mg/L 
NaBz 


0.75 


7.02 


1.63 


68,8 K2S04 


3.70 


3.84 


3.21 


4.30 


3.29 


2.08 


2.54 


0.25 


0.50 


0.42 


58,7 (0.033M) 


3.85 


3.32 


UNSAT. 
SAT. 


8.62 


3.74 


3.94 


8.15 


2.36 


578mg/L 
KN03 


0.28 


7.33 


1.77 


108,12 KN03 


3.33 


3.28 


3.64 


3.38 


2.69 


1.45 


2.08 


0.92 


1.70 


1.27 


98,11 (0.10M) 


3.23 


3.15 


UNSAT. 
SAT. 


8.82 


3.59 


3.30 


8.48 


1.87 


0.01M 
PHENOL 


1.82 


14.9 


3.01 


148.16 Na2S04 


3.29 


3.53 


2.60 


4.59 


3.33 


1.33 


6.26 


1.15 


1.34 


1.14 


138.15 (0.033M) 


3.46 


3.47 


UNSAT. 
SAT. 


14.5 


5.07 


4.57 


26.4 


4.06 


578mg/L 
NaN02 


3.12 


10.7 


3.92 


188,20 LiN03 


2.99 


5.90 


2.43 


4.33 


3.30 


1.38 


2.68 


1.24 


1.71 


1.61 


178,19 (0.10M0 


2.72 


4.43 


UNSAT. 


7.98 


3.63 


4.91 


7.85 


2.15 


0.01M 


2.03 


11.9 


2.36 



DW - deionlzed water 



216 



TABLE 4-B SAT./UNSAT. EFFECT. 
Ag-MINE TAILINGS, 3-inch (i.d.) column; 
AVERAGE (2, columns) 



1.50 kg in each; leachant volume, 350cc. 



f 

Resident time(hr) 
AVG. LEACHANT 

fflNnilfTflNrF KP1 


059 
192 


068 
168 


AVG. 


073 
504 
SAT/UNSAT 


075 
168 


080 
168 


083 088 
504 168 
SAT/UNSAT 

LEACHANT 

DW(ALL) 

12.72 11.68 

16.55 10.85 


LEACHANT 

SLS 

550mg/L 

0.165mmho 

NaBz 

578mg/L 

0.363mmho 

KN03 

O.OIM 

1.507mmho 

PHENOL ■ 

578mg/L 

0.059mmho 

NaN02 

O.OIM 

1 .301mmho 


095 
336 

0.429 
0.521 


104 108 
576 120 
SAT/UNSAT 

LEACHANT 
DW(ALL) 
0.393 0.275 
0.835 0.330 


2&4 (0.10M) 
1&3 12.58mmh 


12.76 
12.75 


13.04 
13.01 


SAT. 

UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 


10.79 
15.54 


13.02 
13.53 


13.42 
13.66 


6&8 (0.033M) 
58,7 7.53mmho 
knot 


7.78 
7.74 


8.03 
7.98 


6.60 
8.92 


8.07 
8.60 


8.59 
8.89 


8.89 
10.60 


6.94 
6.94 


0.567 
0.590 


0.519 
1.105 


0.432 
0.510 


108,12 (O.IOM) 
9&11 11.88mmh 

Na2S04 - 
148.16 (0.033M) 
138,15 6.20mmho 

LiN03 


12.14 
11.97 

6.45 
6.54 


12.56 
12.12 

6.56 
6.58 


9.76 
14.25 


11.89 
12.97 


10.60 
13.31 


12.23 
15.61 


10.94 
10.40 


1.635 
1.720 


1.715 
2.850 


1.411 
1.434 


5.52 
7.67 


6.74 
6.96 


7.11 
6.67 


6.72 

9.46 


5.87 
5.85 


0.106 
0.222 


0.135 
0.546 


0.112 
0.243 


188,20 (O.IOM) 
178.19 8.95mmho 


8.88 
8.86 


8.97 
8.94 


7.81 
10.93 


9.04 
9.41 


0.00 
9.78 


8.92 
11.69 


8.07 
7.38 


1.420 
1.465 


1.476 
2.335 


1.261 
1.209 


LEAD (Pb), mg/L 
28,4 KC1 
18,3 (O.IOM) 


0.22 
0.18 


0.28 
0.39 


SAT. 

UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 

SAT. 

UNSAT. 


0.11 
10.66 


0.22 
0.69 


0.12 
0.21 


0.32 

1.71 


0.14 
0.43 


SLS 
550mg/L 

NaBz 
578mg/L 

KN03 
O.OIM 

PHENOL 
578mg/L 

KN02 
O.OIM 


<.l 
0.14 


<.l 
0.51 


<.l 
0.17 


68,8 K2S04 
58,7 (0.033M) 


0.24 
0.24 


1.31 
1.30 


0.13 
0.48 


0.25 
0.43 


0.13 
0.19 


0.16 
1.39 


0.12 
0.42 


0.09 
0.16 


<.l 
0.55 


<.l 
0.37 


108,12 KN03 
98,11 (O.IOM) 


0.14 
0.14 


0.53 
0.43 


<.l 
0.35 


0.15 
0.32 


<.l 

<.l 


0.11 
1.38 


2.37 
0.23 


0.07 
0.11 


<.l 
5.46 


0.10 
1.88 


148,16 Na2S04 
138,15 (0.033M) 


0.24 
0.27 


0.86 
1.30 


0.17 
3.01 


0.24 
0.64 


0.12 
0.18 


0.19 
2.05 


0.24 
0.24 


0.14 
0.10 


0.10 
0.10 


0.11 
0.14 


188,20 LiN03 
178,19 (O.IOM) 


0.20 
0.19 


n.d. 
n.d. 


0.20 
5.14 


0.29 
0.45 


<.l 
0.20 


0.14 
2.20 


<.l 
0.60 


0.25 
0.25 


0.10 
7.51 


0.11 
1.95 


ZINC (Zn), mg/L 
2&4 KC1 
18,3 (O.IOM) 


0.95 
0.56 


0.75 
0.35 


SAT. 
UNSAT. 

SAT. 

UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 

SAT. 
UNSAT. 


0.56 
5.32 


0.53 
0.96 


0.42 
0.83 


0.63 
12.3 


0.34 
2.60 


SLS 
550mg/L 

NaBz 
578mg/L 

KN03 
O.OIM 

PHENOL 
578mg/L 

KN02 
O.OIM 


0.08 
0.31 


0.39 
3.60 


0.25 
0.86 


68.8 K2S04 
58.7 (0.033M) 


0.54 
0.75 


0.40 
0.43 


0.28 
1.43 


0.27 
0.84 


0.22 
0.66 


0.31 
4.86 


0.15 
1.01 


0.05 
0.18 


0.10 
5.35 


0.08 
1.95 


108,12 KN03 
98,11 (O.IOM) 


0.50 
0.39 


0.44 
0.40 


0.31 
1.86 


0.40 
0.78 


0.31 
0.60 


0.45 
7.20 


0.27 
1.13 


0.28 
0.51 


0.41 
15.4 


0.19 
5.97 


148,16 Na2S04 
138,15 (O.IOM) 


0.44 
0.39 


0.33 
0.85 


0.22 
24.1 


0.23 
2.07 


0.16 
0.82 


0.25 
42.3 


3.65 
1.33 


0.11 
0.61 


0.12 
0.59 


0.06 
0.50 


188,20 L1N03 
178,19 (O.IOM) 


0.39 
0.42 


0.37 
0.44 


0.39 
1.73 


0.57 
0.79 


0.30 
0.70 


0.63 
11.8 


0.36 
1.62 


0.18 
0.98 


0.20 
17.1 


0.24 
3.79 



DW - deionized water 



very similar 
treatment, wh 
effects of mi 
and that atmo 
marily respon 
contamination 
out, however, 
decreases wer 
samples than 



to those observed prior to Both nickel and cadm 

ich would indicate that the detection limits (nickel 

croorganisms were minimal = 0.01) prior to the firs 

spheric oxidation was pri- Concentrations of nickel 

sible for increased leachate (with a Na2SO^ leachant) 

It should be pointed (with LiNC^) for unsatura 

that somewhat greater pH while ranges below detect 

e observed for the untreated Na2S04> to 0.08 ppm (usin 

had been previously seen. observed for those column 



ium were below 
= 0.06, cadmium 
t wet/dry cycle, 
ranged from 0.55 
to 0.12 ppm 
ted columns 
ion (using 
g KC1) were 
s that remained 



217 



saturated. Similar results were also 
seen for cadmium, except values ranged 
from 0.26 (Na 2 S0 4 ) to 0.024 ppm (KNOo) 
for unsaturated columns. All saturated 
columns recorded concentrations below 
detection. 

Acetate Ion Effect 

At the present time, acetic acid 
buffers are used to control the pH of a 
specific leachate. The EP toxicity and 
the TCLP (U.S. EPA 1985) slurry tests 
are two examples where acetic acid is 
used. There have also been suggestions 
involving the mixing of mine tailings 
with other organic matter as a prelude 
to growing a vegetation cover. Since 
acetate ions are known to complex with 
certain metal species (for example, 
lead) , a series of tests were conducted 
to examine the effects of acetate ions on 
the dissolution of metal from the silver 
mill tailings. The results of the ini- 
tial series tests are given in table 5. 
Here, five sets of four columns each 
were leached with leachants of 0.10M. 
NaOAc, pH = 6.03; 0.01M. NaOAc, pH = 6.02; 
0.001& NaOAc, pH = 6.04; 0.10M. KNO3, 
pH = 5.79-control; and 0.10M. NaOAc, 
pH = 6.97 to establish the influence of 
acetate ions. 

All species reported in table 5 
demonstrate increasing metal concen- 
trations with increasing acetate con- 
centrations. Furthermore, the metal 
ion concentrations of 0. 10M_ acetate, 
pH = 6.03, were greater than those of 
O.IOHKNO3, pH = 5.79. This would indi- 
cate that acetate complex formation does 
increase the dissolution of metals within 
this sample of tailings. However, one 
major difficulty remains, i.e., the pre- 
sence of potassium nitrate elevates the 
leachate pH above that of the acetate 
leachate even though the leachant pH is 
lower for the potassium nitrate solution. 
It should also be noted that the metal 
ion concentrations obtained for the 
acetate buffer at pH = 6.97 are somewhat 
greater than those of the potassium 
nitrate control even though the resulting 
leachate pH is greater for the acetate 
buffer. These results would seem to 
indicate that the presence of acetate 
ions will increase the dissolution of 
metals from mine wastes. 

SUMMARY AND CONCLUSIONS 



wastes. It would also appear that the 
chemical nature of the inorganic ion 
within the range of an ionic strength of 
0.1M. has very little effect on the 
concentration of metal species leached 
from the tailings. 

It appears that this sample, which 
is a low-pyrite, high-siderite sample, 
will undergo destructive oxidation 
resulting in decreased leachate pH and 
increased metal dissolution. This oxi- 
dation originates in the unsaturated 
zone and does not appear to involve 
microorganisms. 

The presence of acetate ions has 
a pronounced effect on the dissolution 
of metals from mine waste samples. 
Increased solubility of magnesium, 
manganese, and lead would be expected 
through the formation of acetate com- 
plexes. The dissolution of other 
species, such as zinc, nickel, and 
cadmium, also appears to be increased in 
the presence of acetate ions. Care must 
be exercised regarding actions that could 
lead to acetate production in the proxi- 
mity of mine tailings. 

ACKNOWLEDGEMENTS 

The author wishes to express sincere 
gratitude to C. L. Mardock, Albany 
Research Center, USBM, and J. E. Pahlman, 
Twin Cities Research Center, USBM, for 
their assistance in the mineralogical 
characterization and the destructive 
assay of the tailings samples. A very 
special thanks also goes to Eric Cather, 
Western Field Operations Center, USBM, 
for his microscopic analysis of the sam- 
ples before and after leaching, and for 
many discussions relating to the miner- 
alogy of the systems studied. 

LITERATURE CITED 

Bainbridge, K. L., M. A. Wilkinson, and 
B. M. Mahar. 1980. Evaluation of 
lixiviation of mine waste. Final 
report (USBM contract JO199057) . 
Calspan Corp., Buffalo, NY, 240 pp. 

Ferguson, K. D., and P. M. Erickson. 
1987. Will it generate AMD? An 
overview of methods to predict acid 
mine drainage. Proceedings, Acid 
Mine Drainage Seminar/Workshop 
sponsored by Environment Canada, 
Halifax, Nova Scotia, pp. 215-244. 



Although 1 
from a single m 
study indicate 
on the dissolut 
mine tailings, 
tially explaine 
secondary salt 
activity coeffi 
change between 
tailings. Thus 
with materials 
strength leacha 
dissolution of 



imited to the tailings 
ine, the results of this 
an ionic strength effect 
ion of contaminants from 
This effect may be par- 
d through the standard 
effect (effect on species 
cients) through ion ex- 
the leachant and the 
, treatments of tailings 
that produce high ionic 
te may well enhance the 
specific components in the 



Kuryk, B. A., I. Bodek 
Santhanam. 1985. 
on utility wastes 
experiments, EA-4 
ject 2485-4) . Ar 
Inc., Cambridge, 

Nordstrom, E. K. 1982 
oxidation and the 
tion of secondary 
Ch. in Acid Sulfa 
J. A. Kittrick, D 
L. R. Hossner (ed 
Soc. Amer . , pp. 3 



, and C. J. 

Leaching studies 
: feasibility 
215 (research pro- 
thur D. Little, 
MA, 316 pp. 

Aqueous pyrite 
consequent forma- 
iron minerals, 
te Weathering, 

S. Fanning, and 
s.) . Soil Sci. 
7-56. 



218 



Nordstrom, D. K., E. A. Jenne, and J. W. 
Ball. 1979. Redox equilibria of 
iron in acid mine water. Ch. in 
Chemical Modeling in Aqueous 
Systems, E. A. Jenne (ed.). Amer. 
Chem. Soc. Symp. Series 93, 
pp. 51-80. 



Watzlaf, G. R. 1986. Control of acid 

drainage from mine wastes using bac- 
terial inhibitors. Paper presented 
at 1986 Nat. Meet. Am. Soc. Surface 
Min. and Rec, Jackson, MI, March 
17-20. 

U.S. Environmental Protection Agency. 

1985. Test methods for evaluating 
solid waste. Physical/Chemical 
Methods. SW-846, 2nd ed., rev. 
Environ. Pro. Agen., Washington, DC, 



TABLE 5. ACETATE ION EFFECT 

Ag-MINE TAILINGS 3-inch (i.d.) column; 1, 

AVERAGE (4 columns) « 



50 kg 



pH 



PRIOR 

LEACHANT 

(0W) 



Resident time(hr) 
LEACHANT 

0.10M,KN03,pH=5.79 
0.10M,0Ac,pH=6.03 
0.01M,0Ac,pH=6.02 

0.001M,0AC,pH=6.04 
0.10M,0Ac,pH=6.97 



in each; leachant volume, 350cc. 
110 115 117 119 
144 192 168 168 



7.86 
7.11 
7.82 
7.92 
7.63 



7.72 

7.55 
7.78 
7.91 
7.91 



7.95 
7.61 
7.94 
7.94 
8.20 



8.05 
7.53 
7.77 
7.41 
8.30 



-LEACHANT 

DW 

DM 
DW 
DW 
DW 



3-01 
168 



8.18 
7.63 
8.05 
7.90 
8.38 



3-05 
168 



8.37 
8.20 
8.09 
7.87 
8.52 



CALCIUM (Ca) 
mg/L 



PRIOR 0.10M,KN03,pH=5.79 46.2 34.7 21.2 20.4 DW 15.6 1.05 

LEACHANT 0. 10M,0Ac,pH=6.03 60.3 83.4 65.4 69.6 DW 54.60 4.60 

(DW) 0.01M,0Ac,pH=6.02 16.2 21.3 11.8 12.6 DW 10.3 2.50 

0.001M,0Ac,pH=6.04 19.5 14.8 13.1 13.2 DW 12.0 11.9 

0.10M,0Ac,pH=6.97 60.3 30.3 15.2 14.0 DW 12.8 0.70 



MAGNESIUM (Mg) 
mg/L 



PRIOR 0.10M,KN03,pH=5.79 7.52 3.49 2.27 2.23 DW 1.70 <0.05 

LEACHANT 0.10M,0Ac,pH=6.03 10.2 6.80 6.40 7.15 DW 5.80 0.43 

(DW) 0.01M,0Ac,pH=6.02 1.79 1.97 1.27 1.20 DW 1.05 0.15 

0.001M,0Ac,pH=6.04 2.70 1.80 1.72 1.63 DW 1.50 1.45 

0.10M,0Ac,pH=6.97 13.2 3.58 1.77 1.53 DW 1.53 <0.05 



MANGANESE (Mn) 
mg/L 



PRIOR 0.10M,KN03,pH=5.79 11.3 8.89 2.92 2.22 DW 1.55 0.04 

LEACHANT 0. 10M,0Ac,pH=6.03 18.3 26.8 14.8 13.0 DW 9.43 0.69 

(DW) 0.01M,0Ac,pH=6.02 2.92 2.71 1.62 1.68 DW 1.31 0.36 

0.001M,0Ac,pH=6.04 2.09 1.53 1.20 1.25 DW 1.06 1.28 

0.10M,0Ac,pH=6.97 7.40 4.77 2.17 2.55 DW 2.62 0.04 



NICKEL (Ni) 
mg/L 



PRIOR 0.10M,KN03,pH=5.79 0.12 0.08 0.01 0.03 DW 0.01 <.005 

LEACHANT 0.10M,0Ac,pH=6.03 0.17 0.20 0.20 0.25 DW 0.22 0.01 

(DW) 0.01M,0Ac,pH=6.02 0.03 0.02 <.005 0.01 DW 0.01 <.005 

0.001M.0Ac,pH=6.04 0.04 0.02 0.01 0.02 DW 0.01 0.02 

0.10M,0AC,pH=6.97 0.06 0.02 <.005 0.02 DW 0.01 <.005 



LEAD (Pb) 
mg/L 



PRIOR 0.10M,KN03,pH=5.79 0.28 0.23 0.07 0.36 DW <.03 <.03 

LEACHANT 0. 10M,0Ac,pH=6.03 2.94 6.94 3.40 5.09 DW 1.17 0.11 

(DW) 0.01M,0Ac,pH=6.02 0.18 0.19 0.08 0.19 DW 0.06 0.03 

0.001M,OAc,pH=6.04 0.11 0.08 0.04 0.45 DW <.03 0.03 

0.10M,0Ac,pH=6.97 1.59 1.44 0.58 0.57 DW 0.40 <.03 



0.10M,KN03,pH=5.79 3.74 2.25 0.92 0.91 DW 0.46 0.02 

0.10M,0Ac,pH=6.03 3.91 4.08 5.60 10.42 DW 6.80 0.52 

0.01M,0Ac,pH=6.02 0.95 0.77 0.52 0.55 DW 0.37 0.12 

0.001M,0Ac,pH=6.04 0.40 0.26 0.24 0.26 DW 0.20 0.23 

0.10M.0Ac,pH=6.97 2.99 1.36 0.76 0.86 DW 0.72 0.03 



ZINC (Zn) 




mg/L 


PRIOR 




LEACHANT 




(DW) 



DW - deionized water 



219 



IMPROVED ROCK DURABILITY TESTING TECHNIQUES 
FOR APPALACHIAN VALLEY FILLS 1 



Robert A. Welsh, Jr. and Michael K. Robinson 2 



Abstract. — The integrity and stability of gravity-placed or 
end-dumped excess spoil fills in the Appalachians is partially 
dependent upon placement of durable rock material. Standard 
durability tests do not adequately discriminate between 
nondurable and durable rock. Review of recent rock durability 
research has identified testing procedures and classification 
systems that may be more applicable to excess spoil fills and 
the surface mining process. A geotechnical testing program, 
incorporating the transfer of this recent technology, has been 
designed by the U.S. Office of Surface Mining Reclamation and 
Enforcement to allow better prediction of rock durability for a 
range of overburden materials. An array of rapid, inexpensive 
rock competency tests are being compared to determine which 
tests or combination of tests give correlative results which 
allow accurate prediction of rock durability. Preliminary 
results from 18 rock samples collected from minesites in 
Kentucky, Virginia, and West Virginia suggest that tests which 
measure rock swelling upon water immersion are particularly 
valuable in predicting the behavior of marginally durable 
shales. Simple swell tests, when combined with rock strength 
tests, seem to provide the most efficient discrimination 
between rocks of similar appearance, but widely varying 
durability. An additional 84 rock samples collected from 43 
minesites in Virginia, West Virginia, Kentucky, and Tennessee 
are being analyzed to further refine the testing techniques and 
to investigate the primary mechanisms contributing to slaking. 






INTRODUCTION 

Excess spoil consists of overburden (soil and 
rock excavated during the mining operation) not 
needed to reclaim the disturbed area to the 
approximate original contour. Before the Surface 



''Paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the 
Interior (Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), 
April 17-22, 1988, Pittsburgh, PA. 

2 Robert A. Welsh, Jr. is Geologist, and 
Michael K. Robinson is Supervisory Physical 
Scientist, Eastern Field Operations, USDI Office of 
Surface Mining Reclamation and Enforcement, 
Pittsburgh, PA. 



Mining and Reclamation Act of 1977 (SMCRA), excess 
spoil structures were constructed with minimal 
engineering guidance. 

Often these structures were placed at 
locations selected strictly to optimize mining 
operations. Little thought was given to potential 
environmental consequences or safety hazards. 
Since the passage of SMCRA, there has been an 
increase in the engineering effort directed toward 
design and construction of excess spoil disposal 
areas. 

In general, methods of placement for excess 
spoil include: (a) the lift type construction 
method; (b) the head-of-hollow fill method; and 
(c) the durable rock (gravity) fill method. 



220 



In the lift method, excess spoil is usually 
deposited in uniform, horizontal lifts of 4 ft or 
less and compacted to achieve the desired density. 
Prior to placement of the spoil in this type of 
fill, the foundation must be prepared and 
underdrains installed. According to U.S. Office of 
Surface Mining Reclamation and Enforcement (OSMRE) 
regulations at 30 CFR Section 817.71(f)(3), the 
rock comprising the underdrain must be durable 
(rock that will not slake in water nor degrade to 
soil material); non-acid or toxic forming; and free 
of coal, clay or other non-durable material. 

An alternative method for excess spoil 
disposal involves the placement of spoil in lifts 
at the upper reaches of a watershed. This "head- 
of-hollow fill" method originated in West Virginia 
in the early 1970's, and combines the lift- 
placement technique and a durable rock chimney 
drain in the center of the fill. The "rock core 
chimney drain" results from physical segregation of 
larger rock during spreading of spoil material and 
lift compaction. All surface and subsurface 
drainage is to be controlled by this rock core, in 
order to prevent elevation of the phreatic surface 
within the fill mass. This type of fill must be 
placed where the surface drainage entering the core 
is minimized to prevent a decrease in permeability 
due to clogging of the rock core by fine 
particles. 

The durable rock fill method consists of 
dumping spoil to its angle of repose into valleys 
in a single high lift or several smaller lifts. In 
existing fills, the lifts range between 50 to over 
400 ft in thickness. The front face of the fill is 
then graded to develop a terraced fill 
configuration. The material forming the rock fill 
is generally made up of angular blast rock. 
According to 30 CFR Section 816.73, the durable 
rock fill method can only be used if durable rock 
overburden is present and comprises at least 80 
percent by unit volume of the fill. No designed 
underdrain is required for this type of fill, in as 
much as the gravity segregation which occurs upon 
dumping forms a highly permeable zone of large- 
sized durable rock in the lower one-third of the 
fill. 

The successful performance of excess spoil 
structures is directly related to the durability of 
the rock in the fill mass and underdrains. As D.R. 
Casagrande noted in a public hearing on OSMRE 
proposed rules concerning spoil in fills: 



long term, a durable rock fill should behave more 
as a rock mass than as a soil mass. A rock mass is 
inherently more stable than a soil mass of similar 
volume because rock has much greater load-carrying 
capacity and resistance to movement or 
consolidation than soil. Durable rock fill 
material has this greater strength because of high 
intergranular friction and greater resistance to 
shear stress. Nondurable rock will degrade into 
soil-sized particles as a result of overburden 
pressure and moisture absorption, and the drainage 
system provided by the void space between the rocks 
may become clogged. The clogging may cause excess 
pore water pressures to develop that will cause a 
decrease in the shear strength of the fill 
material. This decrease in shear strength can 
cause the failure of the excess spoil structure. 
Therefore, the correct assessment of the durability 
of the rock is a critical design factor. 

The OSMRE recognizes the need for a suitable 
rock durability standard. The objective is to 
select a rapid, inexpensive durability testing 
standard which will clearly differentiate between 
durable and nondurable materials, model the effects 
of surface mining conditions on rock materials, and 
allow assurance of the long-term stability of 
properly designed fill structures. 

Durability classification systems that involve 
more than two tests may be uneconomical and are 
subject to the accumulative effect of mechanical 
and human errors during testing. Franklin (1970) 
and Bieniawski (1974) consider the following as 
necessary prerequisites for any rock classification 
system employed on a routine basis: 

1 . System should be based on measurable 
parameters determinable by relevant tests 
performed quickly and inexpensively in the 
field; 

2. System should involve only rapid testing 
techniques due to the potential for large 
numbers of routine samples; 

3. Testing techniques should be simple enough 
to be carried out by semiskilled field and 
laboratory staff; and, 

4. The range of test result values should 
allow for a sufficient power of 
discrimination when applied to the various 
test samples. 



"Spoil materials range from hard rocks to clay 
shales and even soft clay. The range of 
engineering properties of such materials is 
enormous. Therefore the proposed rules must 
be sufficiently conservative to also include 
the properties of the weakest materials." 
(Casagrande, 1978). 

Durable rock is defined in Federal regulations 
at 30 CFR Section 816.73(b) as rock which does not 
slake in water and will not degrade to soil 
material. A rule-of-thumb used by regulatory 
authorities is that durable rock achieve an as- 
tested slake durability index (I D ) of at least 90. 
The intent of the durability standard is to 
selectively obtain rock that can withstand surface 
mining conditions including blasting, handling, 
compaction, and weathering without significant 
degradation. The key concern is that, over the 



MATERIALS AND METHODS 

Samples 

The rock samples tested in this study were 
collected from recently blasted highwalls of 
surface mines in Kentucky, Virginia, and West 
Virginia. Eighteen grab samples of freshly blasted 
rock weighing approximately 100 lb were collected 
at each site. To facilitate coring during 
laboratory sample preparation, large, competent 
blocks of rock were selected. Detailed 
descriptions of geologic properties and mines ite 
conditions relating to rock durability were made on 
site, and photographs were taken of the sampled 
highwall. A dilute hydrochloric acid effervescence 
(fizz) test for calcareous cementing agents was 
performed for each sample. 



221 



Early to Middle Pennsylvania-age sandstones, 
shales, and mudstones were sampled for this study. 
These include overburden and interburden rocks of 
the Hazard and Peach Orchard Coal zones of the 
Breathitt Formation in Kentucky; Standiford, 
Taggart Marker, and Low Splint Coal zones of the 
Wise Formation in Virginia; and, Freeport and 
Kittanning Coal zones of the Allegheny Formation, 
and the Coalburg Coal zone of the Kanawha Formation 
in West Virginia. 



Rock analyses 

Geologic materials removed from their in situ 
environment during the surface mining of coal 
exhibit changes in physical integrity. Such 
changes are caused by physical and chemical 
mechanisms induced by variations in moisture and 
stress regimes. The rock in fills has been 
subjected to blasting, handling, compaction, and 
weathering. Generally speaking, a sedimentary rock 
that can withstand these processes without 
significant changes in its original structure can 
be classified as a durable rock. 

When selecting durable rock for fills, one 
should choose a single test or a combination of 
tests that best simulate surface mining conditions 
(Robinson and Ventura 1983). For this study, 
recent research on rock durability has been 
reviewed, and testing techniques were selected for 
application to surface mining rock fills. In order 
to establish which test or combination of tests 
best serves as an indicator of sedimentary rock 
durability, a laboratory testing program was 
designed to simulate the moisture changes and 
stress regimes that a sedimentary rock undergoes 
during the processes of excavation and placement in 
excess spoil fills. After conducting a wide 
variety of tests, including Modified Los Angeles 
Abrasion testg, Modified Slake Durability Index 
tests, and US Army Corps of Engineers (COE) 
Accelerated Weathering tests, among others, a 
simplified testing protocol evolved. 



c) Atterberg Limits Testing is an indicator 
of the type of clay minerals in the rock 
and their plasticity. This involves 
measuring the liquid limit, plastic limit, 
and plasticity index of the fine grained 
(minus #40 sieve size) fraction of the 
rock material. 

d) Swell Testing indicates the slaking stress 
that affects rock when it is removed from 
its in situ environment in the overburden. 
This is done by measuring the volume 
expansion of the cored rock normal to 
bedding upon immersion in water for a 
period of 24 hours after oven drying to 
105 degrees C. A dial gage is used to 
measure any dilatancy. Swelling strains 
are probably the result of expansion due 
to air breakage along interconnected 
voids such as microcracks in the rock 
(Olivier 1979). 

e) Jar Slake (Soak) Tests were run to provide 
a qualitative measure of rock behavior 
after immersion in water for a 24-hour 
period. The sample is immersed in a 
beaker, and observations of any 
disaggregation are made. As Andrews et 
al. (1980) state, "...this test might be 
most closely related to spoil materials 
located at depth and within a constant 
humidity or totally saturated 
environment."; therefore these tests are 
particularly relevant to the durability of 
underdrain rock in valley fills. 

The above tests were conducted by the COE Ohio 
River Division Geotechnical Engineering Laboratory 
under the testing protocol designed by OSMRE. In 
this discussion, test results for only a limited 
number of sedimentary rock samples are reported. 
Thus, any conclusions presented herein should be 
considered as preliminary until results are 
available for a wider range of samples currently 
being tested. 



The program presently includes the following 



tests: 



RESULTS AND DISCUSSION 



a) Slake Durability Testing , which is 
currently accepted by OSMRE, includes 
minor abrasion effects and saturation- 
desiccation stresses. As defined by 
Chandra (1970) and Franklin and Chandra 
(1972), oven-dried samples of rock are 
placed in a wire mesh drum partially 
immersed in water. The drum is rotated at 
20 rpm for approximately 10 minutes; the 
sample is then removed, dried, and run 
through a second cycle. 

b) Uniaxial Unconfined Compressive Strength 
Tests simulate loading stresses in a fill. 
They are run on rock cores loaded in the 
direction normal to bedding. Loading is 
applied to the point of breakage, defined 
as the maximum stress; constant loading 
rate of 8,000 lb/min was used. 



Data from the laboratory testing program are 
listed in tables 1 and 2. The values obtained in 
the laboratory for parameters including the 
swelling strain, slake durability index, unconfined 

Table 1. — Atterberg limits data and Deere and 
Gamble (1971) plasticity classifications 
for sampled Appalachian shales. 



Sample 


Liquid 


Plastic 


Plasticity 


Plasticity 


Number 


Limit 


Limit 


Index 


Rating 




(%) 


(%) 


{%) 




RR-1 


30 


21 


9 


Low 


CB-3 


33 


22 


11 


Medium 


KS-1 


30 


22 


8 


Low 


KS-2 


32 


22 


10 


Low/Medium 


AR-1 


28 


21 


7 


Low 


AR-2 


35 


27 


8 


Low 


WR-2 


36 


26 


10 


Low/Medium 


IF-3 


35 


24 


11 


Medium 


RR-1 a 


37 


22 


15 


Medium 



222 



Table 2. — Swelling coefficients, compressive strengths, slake durability 
indices, and Franklin and Chandra (1972) and Olivier (1979) 
classifications for Appalachian rock samples. 













Durability 






Swelling 


Compressive 


Slake 


Classification 




Franklin/ 


Olivier 


Sample 


Rock 


Ratio 


Strength 


Durability 


Chandra 


(1979) 


Number 


Type 


(DL/L) 


(lb/in 2 ) 


Index (J) 


(1972) 




RR-1 


Shale 


0.0085 


11 ,880 


98.5 


V.high 


M.poor 


CB-3 


Shale 


0.0486 


930 


94.8 


M.high 


V . poor 


KS-1 


Shale 


0.0125 


2,850 


97.2 


High 


V.poor 


KS-2 


Mudstone 


0.0299 


2,210 


94.9 


M.high 


V.poor 


AR-1 


Shale 


0.0020 


3,060 


95.9 


High 


M.poor 


AR-2 


Shale 


0.0116 


1,670 


96.4 


High 


V . poor 


WR-2 


Mudstone 


0.0142 


- - 


29.6 


V. low 


V.poor 


IF-3 


Shale 


0.0280 


180 


92.1 


M.high 


V . poor 


RR-1 a 


Shale 


0.0010 


3,080 


95.9 


High 


Fair 


RR-2 


Sandstone 


- - 


14,280 


98.3 


V.high 


Excellent 


CB-1 


Sandstone 


0.0018 


6,670 


92.5 


M.high 


Excellent 


CB-2 


Sandstone 


- - 


6,050 


84.0 


Medium 


Excellent 


KS-1 a 


Sandstone 


0.0004 


3,480 


95.7 


High 


Excellent 


PH-1 


Sandstone 


0.0002 


4,200 


89.1 


M.high 


Excellent 


AR-3 


Sandstone 


0.0003 


6,630 


97.9 


High 


Excellent 


WR-1 


Sandstone 


- - 


5,060 


96.8 


High 


Excellent 


IF-1 


Sandstone 


0.0008 


5,150 


94.3 


M.high 


Good 


CB-1 a 


Sandstone 


0.0006 


3,660 


96.7 


High 


Good 



- - indicates value too low for accurate measurement. 

M. = moderately. 

V. = very. 

DL/L = ratio of change in sample length to original sample length. 



uniaxial compressive strength, and Atterberg limits 
of the rocks, wherever applicable, were used to 
classify their durability under the classification 
systems developed by Franklin and Chandra (1972), 
Deere and Gamble (1971), and Olivier (1979). None 
of the rock samples had significant calcareous 
cement, based on negative fizz test results. 

Durability classification of the sampled rock 
is shown graphically in figures 1 through 3, using 
single and multiple index classification systems. 
The single index classification system developed by 
Franklin and Chandra (1972) uses only the slake 
durability index test data to assess durability. 
Slake durability of the rock is assessed by an 
index, Ip, defined as the percentage retention 
measured by dry weight after two cycles of testing 
(fig. 1). This test was adopted by 0SMRE as the 
accepted standard, and thus has been widely used 
among the coal industry for the selection of 
durable rock. Figure 1 indicates that, for the 
rocks sampled in this study, the slake durability 
test lacks sufficient discrimination to reveal 
durability differences between sedimentary rocks as 
disparate as sandstone, shale, and mudstone. 
Furthermore, the overwhelming majority of sampled 
rocks are uniformly ranked as highly durable. 
Several authors have also reported similar 
problems in the use of the slake durability index 
as applied to geotechnical projects such as highway 
or tunnel construction (Duncan et al. 1968, Noble 
1977, Olivier 1979, Richardson 1985). Published 
difficulties with the test include: 



1. During testing, some of the more plastic 
shales may form mudballs, thus rendering 
falsely high I D values (Richardson 1985). 

2. The test does not differentiate very well 
between shales (Noble 1977). 

3. The test is insensitive to shales which 
slake into small chips which are larger 
than the #10 sieve (2 mm) size of the 
openings in the hardware cloth forming the 
testing drums (Duncan et al. 1968, 
Olivier 1979). This effect was 
duplicated in shale samples tested in the 
present study, where high I D s of over 90 
were attained, yet the shale samples 
disintegrated into masses of small chips 
in the soak test (fig. 4). Breakdown of 
rock into such small soil-like particles 
in fills may lead to soil-like behavior 
under the large load factors developed in 
the fill mass, and fill failure may 
result. 

Additional concerns specific to surface mining 
were raised by Welsh et al. (1985): 

1 . The test fails to subject rock samples to 
the types of physical stresses common to 
surface mining conditions (impact, heavy 
abrasion, saturation and desiccation, 
compaction, etc.). 



223 



Very Low 




2 Cycle 
Classification 



1 Cycle 
Classification 



99 100 
'High' 



Slaking Durability Index (% Retained) 

Legend 

O Sandstone 
• Shale 
A Mudstone 

Figure 1.— Classification of Appalachian rock samples under the Slake Durability Index System of 
Franklin and Chandra (1972). 



2. The index does not assess properties of 
rock samples indicative of rock-like or 
soil-like behavior. 

3. Samples classified as durable in the 
laboratory exhibit soil-like behavior in 
the mining process. 

Recognizing these problems in using the slake 
durability test exclusively as a single index 
measure of rock durability, other classification 
systems incorporating additional geotechnical tests 
were compared to the slake durability test 
standard. Dual-index graphical classification 
systems developed by Deere and Gamble (1971) and 
Olivier (1979) were utilized to provide a basis for 
correlation of durability measures. 

The Deere and Gamble (1971) durability 
classification system is based on the two-cycle 
slaking durability and the plasticity index of 
sedimentary rocks. The data from the present study 
indicate a 3lightly better resolution between 
shale, mudstone, and sandstone durability as 
classified under the Deere and Gamble (1971) 
system, compared to slake durability testing alone 
(fig. 2). Sandstones, since they exhibit 
negligible plasticity, plot at the base of the 
chart; while shales and mudstones are distributed 
across the chart by their plasticity values. 
Gamble (1971) and Deere and Gamble (1971) found 



from laboratory testing that the plasticity of 
sedimentary rocks is inversely related to their 
slake durability. Therefore, higher plasticity 
values would decrease the rock durability. 
However, the Deere and Gamble (1971) system makes 
no attempt to adjust slake durability test rankings 
based on plasticity values. This classification 
system suffers from the same problems as the 
Franklin and Chandra (1972) test because it 
incorporates slake durability test data. 

Olivier (1979) and Duncan (1969) recorded 
sedimentary rock behavior when it is removed from 
the in situ environment. Rocks swell and 
disintegrate as a result of stress relief. This 
swelling increases as rock absorbs moisture. 
Olivier (1979), also found that as swelling 
increases, the uniaxial compressive strength of the 
rock decreases. The resulting rock durability 
classification system involves the measurement of 
two rock properties. The first parameter is the 
magnitude of rock swelling after a dried sample is 
immersed in water. The second is the uniaxial 
compressive strength. Broad categories of 
"geodurability" ranging from very poor to excellent 
were assigned by Olivier (1979) based on ratios of 
compressive strength to swelling coefficient. 

Classification of the rocks sampled in the 
present study using the Olivier (1979) system shows 
good discrimination between rock types and even 
between rocks of the same type (fig. 3). 



224 



i 

X 

« 

73 

C 

I 



100- 
90- 

80-- 

70-- 

60-- 

50-- 



£ 40- 

CO 
CO 



30-- 



20-- 



10- 



Very Low S.D. 



Low S.D. 



Medium / r- 
Medium S.D . , ., High, ,/,,/( s.D. 



High S.D. 
Very High 



S.D. 



O 000 3800 



High P.I. 



Medium P.I. 



Low P.I. 



10 20 30 40 50 60 70 80 85 90 95 100 

S.D.: 2 Cycle Classification (% Retained) 

Legend 
o Sandstone 
• Shale 
A Mudstone 



Figure 2. —Classification of Appalachian samples under the Deere and Gamble 
(1971) System. 



Sandstones plotted in the good to excellent 
geodurability class, while shales ranged from fair 
to very poor. Muds tones ranked as very poor. The 
shales plotted into two subpopulations, grouped as 
moderately poor-fair, and very poor. Such 
results, if substantiated by further testing of a 
variety of rock, could be used to distinguish 
between tougher, more durable shales, and shales 
and mudstones which should be considered as 
nondurable. 

Results from the soak tests support the 
Olivier geodurability assessments. Shales with 
slake durability indices of at least 90 (high 
durability) were rated as poor in geodurability, 
and disaggregated into fine, soil-like particles ir 
the soak test (fig. M). Therefore it appears that 
for the samples tested, the slake durability 
testing was not rigorous enough to accurately 
assess rock durability. 



SUMMARY AND CONCLUSIONS 

Comparison of the three classification systems 
indicates that all the systems correlated well in 
classifying the sampled sandstones as rocks with 
generally high durability. For shale and mudstone 
durability, however, the classification systems 
varied in usefulness. The Franklin and Chandra 
(1972) slake durability index and the Deere and 
Gamble (1971) classification system did not 
effectively distinguish between rock types of 
different durability. The reliance of these 
classification systems on the slake durability 
index test, which has inherent problems identified 
earlier, is the reason for this failure. 



225 



Very Low Strength 



Low . . Medium. 



Very 
Strength Stren 



r en jth 




.0001 



J_L£. 
3 4 5 10 20 30 40 50 100 

Uniaxial Compressive Strength <%, (MPa) 
Legend 



o Sandstone 
• Shale 
A Mudstone 



if: Sandstone with Co <. 0001 
■^f Mudstone with <y -c 1 



Figure 3. — Classification of Appalachian rock samples under the Olivier (1979) 
Geodurability System. 



The Olivier (1979) classification system shows 
promise as a more accurate measure of rock 
durability, particularly for rock materials of 
marginal durability, such as shale. Correlation 
with soak test results indicates that the swell and 
compressive strength tests measure factors which 
are pertinent to the original concerns regarding 
rock durability in surface mining valley fills. 

Further testing will be performed at COE 
laboratories to assess the reproducibility of 
results utilizing a testing protocol including 



swell, compressive strength, and soak tests on 84 
additional rock samples from 43 minesites in 
Kentucky, Tennessee, Virginia, and West Virginia, 
and to compare these assessments with those from 
the slake durability index test. Point load 
testing results will be compared to uniaxial 
compressive strength testing, for possible 
interchangability of these rock strength tests. 
The geodurability rankings in the Olivier (1979) 
system will be modified to reflect experience 
gained through the testing of Appalachian rock 
material to better assess the durability of rock 
placed in excess spoil fills. 



226 




Sample K S-2 before Soak Test. 




Sample K S-2 after 24-hour Soak. 

Figure 4. — Qualitative scak test result for 
Appalachian shale sample. 



LITERATURE CITED 

Andrews, D. E., Withiam, J. L. , Perry, E. F., and 
H. L. Crouse. 1980. Environmental effects of 
slaking of surface mine spoils: eastern and 
central United States. Bureau of Mines, 
U.S. Department of the Interior, Denver! CO. 
Final Report, 247 pp. 

Bieniawski, Z. T. 1974. Geomechanics 

classification of rock masses and its 
application in tunneling. Proceedings of the 
3rd International Congress on Rock Mechanics. 
Denver, CO. Vol II, pp. 27-32. 

Casagrande, D.R. 1978. Presentation at Public 
Hearings. October 26, 1978. Submitted as 
written comments on the letterhead of 
Casagrande Consultants, October 27, 1978. 3 
pp. with 4 page attachment. 



Chandra, R. 1970. Slake durability tests for 
rocks. Master's Thesis, Imperial College, 
London University. London., 55 pp. 

Deere, D. U., and J.C. Gamble. 1971. Durability- 
plasticity classification of shales and 
indurated clay. Proceedings of the 22nd 
Annual Highway Geological Symposium, Norman, 
OK. pp. 37-52. 



Duncan, N. 1969 
mechanics. 



Engineering geology and Rock 
252 pp. Leonard Hill, London. 



Duncan, N., Dunne, M.H., and S. Petty. 1968. 

Swelling characteristics of rocks. Water 
Power. May 1968:185-192. 

Franklin, J. A. 1970. Observations and tests for 
engineering description and mapping of rocks. 
Proceedings of 2nd International Congress on 
Rock Mechanics, Belgrade, Vol. 1, pp 1-3- 

Franklin, J. A., and R. Chandra. 1972. The slake 
durability test. Int. J. of Rock Mech. and 
Min. Sci. 9:325-341 . 

Gamble, J. C. 1971. Durability-plasticity 
classification of shales and other 
argillaceous rocks. Ph. D. Thesis, University 
of Illinois, 161 pp. 

Noble, D.F. 1977. Accelerated weathering of tough 
shales, final report. Virginia Highway and 
Transportation Research Council. 
Charlottesville, VA. No. VHTRC 78-R20. 38 pp. 

Olivier, H. J. 1979. A new engineering-geological 
rock durability classification. Engr. 
Geol. 14: pp. 255-279. 

Richardson, D.N. 1985. Relative durability of 
shale - a suggested rating system. 
Proceedings of the 36th Annual Highway Geology 
Symposium. Clarksville, IN. pp. 105-138. 

Robinson, M. K. , and J.D. Ventura. 1983. Disposal 
of excess spoil: durable rock fills. 
Symposium on Surface Mining, Hydrology, 
Sedimentology and Reclamation, Lexington, 
Kentucky, pp. 179-187. 

United States Code of Federal Regulations. 1987. 
Title 30, Mineral Resources, Chapter VII, 
Enforcement, Department of the Interior. 
Section 817.71, pp. 277-278. 

Welsh, R.A., Robinson, M.K., and L.E. Vallejo. 
1986. Evaluation of durability testing 
techniques for rock underdrain material used 
in Appalachian surface coal mining valley 
fills. Proceedings of the International 
Symposium on Flow- Through Rock Drains. 
Cranbrook, BC. pp. 83-93- 



227 



ACID MINE DRAINAGE RESEARCH: HYDROLOGY'S CRITICAL ROLE AND UNIFYING THEME 



Frank T. Caruccio 1 



Abstract. — With some exceptions, in innovative 
treatment techniques many of the concepts and 
theories regarding acid mine drainage were 
postulated over 3 5 years ago. And much of the 
mine drainage related research currently 
performed should be integrated by a unifying 
concept and directed toward the solution of a 
principal objective - namely, the in-situ 
abatement of acid mine drainage. A central 
element which can provide a vehicle for unifying 
research directed toward this end and which is 
least understood and studied, is the hydrology of 
a backfill. A discussion is presented to justify 
soliciting proposals directed primarily toward 
and better understanding of the nature, 
occurrence and movement of water within 
spoil/backfill material. 



Knowledge of the occurrence, 
distribution and description of water 
movement in a backfill, collectively 
defined by the hydrology, is critical to 
an understanding of the chemical evolution 
of water as it recharges the backfill and 
eventually emerges as a seep. The quality 
of the seep produced determines whether or 
not a bond is released, impacts the 
overall economics of the mining operation 
and affects the permitting decisions of 
future mines in that area. To date, few 
studies have addressed the hydrology of 
backfilled sites, primarily, I suspect, 
because of the difficulty of obtaining a 
hydrologic data base that requires 
drilling of the backfill material and 
properly installed wells and piezometers. 
Other than the ground water installations 
required to monitor quality, there is 
little incentive to drill costly 
monitoring wells in a backfill. Of the 
mine drainage issues, the hydrology of the 
system, although the most important part 
of the problem, remains the biggest 
unknown. If there is little incentive to 
drill and thoroughly explore an acid 
producing reclaimed site to determine what 
went wrong, imagine the incentive to 



1 Frank T. Caruccio is Professor of 
Geological Sciences, Department of 
Geological Sciences, University of South 
Carolina, Columbia, SC 29208. 



hydrologically explore a non-acid site to 
find out what was done right! Let's look 
first at the acid side of the issue. 

From an historical perspective, acid 
mine drainage has been a persistent 
problem, plaguing and limiting coal 
development as a viable energy base. 
Although simple to explain, the problem 
has yet to be solved on a level whereby 
regulatory agencies may feel comfortable 
in granting permits to mine certain seams 
in certain areas. Indeed, during the last 
30 years substantial progress has been 
made in the prediction methodology, 
understanding the nature of the problem 
and in treatment technology (particularly 
in reclamation of landscapes and treatment 
of acidic mine drainages) . Yet, in those 
problem areas, acid waters continue to be 
produced and although recent mine plans 
are designed to minimize or contain 
underground the acid loads, the mining and 
reclamation strategy necessary to 
completely eliminate acid mine drainage 
continues to elude us. 

Further, the cost of treating and 
controlling acid mine drainage has been 
estimated to be in excess of 5 billion 
dollars (1970 dollars)- one wonders how 
this entire scenario can be economically 
viable! The severance taxes and 
reclamation funds instituted by the Mining 
Act, that are currently being returned to 
the states, are beneficial but the 



228 



reclamation efforts are usually directed 
toward alleviating mine related problems 
which remotely address acid water 
abatement. More often than not, 
considerable efforts are directed toward 
the revegetation of an abandoned mine 
site, with peripheral attempts toward 
water quality amelioration; in all 
fairness probably due to a lack of a 
demonstrated effective in-situ treatment 
rather than intent. 

During my 25 years or so in acid mine 
drainage research I initially directed my 
efforts toward an understanding and 
prediction of acid mine water. Over the 
last ten years my efforts have been 
directed toward the mitigation of acid 
waters through the in-situ manipulation of 
the hydrology and geochemistry of 
infiltrating or recharge waters. As part 
of this overall effort, several ancillary 
studies developed with regard to 
prediction technology and the hydrology of 
a backfill. And over the past 15 years or 
so, through several committee 
appointments, I was able to periodically 
review the state of the art of acid mine 
drainage research, nationally as well as 
internationally. Based on what you are 
familiar with, consider the following 
concepts and their applications to the 
abatement of acid mine drainage: 

1. The acidity of acid mine drainage 
originates not from free acid but from the 
hydrolysis of certain iron sulfates ( by 
products of pyrite oxidation) . 

2 . The occurrence of the problem is 
geographically limited to certain parts of 
the bituminous coal field. 

3 . Concentrations of acidity vary with 
conditions of flow to produce variations 
in loads. 

4 . The primary source of acidity is 
pyrite or marcasite in the roof shales, 
binders and rider seams. 

5. Acid formation is associated with 
the concurrent growth of iron-oxidizing 
bacteria and the bacteria serve as 
catalysts. 

6. The key factors in the formation of 
acid mine drainage are the presence of 
finely divided pyrite which is exposed to 
the atmosphere and the oxidation of 
ferrous by bacteria. 

7 . The problem is related to the 
occurrence of sulfates of iron, aluminum 
and manganese. 

Further, recommended research on 
control programs might be aimed at: (a) 
the control of the disposition of the iron 
sulfides, (b) prevention of contact of 
iron disulfide with air and water, (c) 
reduction of bacterial activity and (d) 
removal of the objectionable impurities 
from acid mine water. 



Sound familiar? ALL of the above were 
gleaned from literature dated no recent 
than 1954! The classic works of S.A. 
Braley, L.V. Carpenter, R.L. Starkey, K.L. 
Temple and W.A. Koehler, to name a few, 
and my apologies to the other giants in 
the field who were not mentioned, go back 
to the early 1930' s in publication dates. 
Other than the few major breakthroughs, 
primarily in treatment, what is the stage 
of our progress to date and since then, 
how have the overall research efforts been 
structured to build upon these concepts 
and observations reported over 35 years 
ago? 

A review of the recent literature in 
acid mine drainage research shows that 
little has been added to the basic pool of 
knowledge central to the mitigation of 
acid formation at the pyrite site. 
Several breakthroughs have occurred in the 
treatment technology, notably, the 
wetlands treatment of acid waters, the 
surfactant bactericides developed to 
control the iron oxidizing bacterial 
activity and the large scale treatment 
facilities designed to handle large 
volumes of acidic water. However, how 
much closer are we toward solving the 
problem? It seems to me that the acid mine 
drainage problem which in itself is 
overwhelming, has been dissected into 
minute parcels of research, for funding 
expediency and/or clarity of thought, 
which have been taken independently away 
from the central objective related toward 
the solution of the problem. Today, with 
some notable exceptions, very little 
research is integrated into a unifying 
conceptual theme. Commonly, the results 

are reported to answer the objectives of 
the study and little is done to take the 
data and conclusions from that realm into 
the real world through the synthesis of 
ideas or integration into the other facets 
of the acid mine drainage problem. For 
example, the worth or value of a study 
that shows a chemical aspect of the acid 
problem would be greatly enhanced if the 
results of the study can be integrated 
into the hydrology of the site and further 
developed toward field application. 
Unfortunately, the research has been so 
finely refined that the trees become 
identified as not being part of the 
forest. 

As a first step toward addressing the 
solution of the acid mine drainage problem 
through a perpetual and effective in-situ 
abatement technique let me suggest that 
future research in acid mine drainage 
abatement studies be structured to 
interrelate two or more aspects of the 
natural system and restrict funding to 
those projects that can demonstrate that 
the final results can be easily integrated 
into an ever expanding data base. The 
pyramid concept should be introduced with 
the keystone being the solution to the 
problem and supporting research 
constituting the building stones and 
blocks that establish the base, but at the 



229 



same time converge toward the central 
objective. It seems that today we 
continue to lay a broader and broader base 
with no sense of convergence occurring 
within the immediate future. 

In soliciting proposals, funding 
agencies commonly call for studies within 
extremely broad categories the are related 
to the problem, but which do not provide 
convergence or direction toward addressing 
a central objective. Consider the RFP 
published in the September 18, 1987 
Commerce Business Daily, soliciting 
research proposals "... to provide 
practical solutions to selected problems 
of abandoned coal mines. Proposed 
research must result in the development of 
innovative technologies, have wide 
regional application and be completed in a 
short time frame." The request then goes 
on to say tnat they are particularly 
interested in the areas of "...water 
quality, specifically abatement of acid 
mine drainage from a) underground mines, 
and b) refuse piles..." 

Understandably the request is 
structured in the broadest description to 
encourage novel, innovative and unique 
approaches that may not as yet been 
conceptualized. I agree with this 
strategy. But at the same time could not 
part of the solicitation be line item 
specific, calling for directed research 
leading toward a specified objective. 
This may invoke a response from an 
expertise or talent, though not familiar 
with the overall problem of acid mine 
drainage, who may become involved in the 
research through directed efforts. For 
example, one of the key elements in the 
problem being discussed is the occurrence 
of water within a refuse pile or backfill 
and the critical role that it plays in the 
mass transfer and production of acidity 
and alkalinity. How is water recharged to 
the spoil, what is the distribution of the 
pores, what rock type generates the 
greatest porosity and provides the 
greatest alkaline/acid potential, what is 
the residence time of the water recharging 
the spoil, and how are the discharges 
controlled at the toe or outflow to 
produce seeps? Are the seeps an admixture 
of multi-channel flows or large reservoir 
bleeds? Each of these is a major unknown 
and must be addressed. In addition to the 
solicitation described above could not 
these research needs be specifically 
identified and targeted for proposals? A 
master plan must be organized that 
contains a convergence of studies and 
research efforts directed toward the 
solution of the problem. 

I am a firm believer in the value of 
basic research. Certainly, it is through 
this avenue that data bases are developed, 
which, in turn, are synthesized into 
hypotheses and become foundations for 
applied and pragmatic research. But, 
realistically, how much closer are we 
toward solving the problem than we were 
twenty five years ago. Over the years, 



and especially within the last ten years, 
a substantial amount of time, effort and 
money has been spent on the acid mine 
drainage problem. From my perspective, 
the data and conceptual base has been 
expanded, tremendously. Developments in 
predictive technology have provided 
insight into the nature and character of 
sulfide reactivity, reclamation efforts 
have provided insight into the mature of 
the pore gases and their evolution giving 
us a better understanding of the 
hydrogeochemistry of the problem. 
Treatment technology advances have given 
us a variety of mitigative applications 
that may or may not be site specific, but 
certainly are restricted to specific 
settings within the backfill or spoil pile 
(i.e. in order to be effective, phosphate 
amendments must be juxtaposed to reactive 
sulfide sites-consequently are limited in 
application to active or reconstructed 
sites) . And numerous studies 
characterizing the physical aspects and 
geometries of the problem sites have 
yielded broad insights into the chemistry, 
hydrology and biology of the problem. The 
time has come to reexamine the broad base 
of completed studies that are laterally 
distributed and to recognize a common 
element upon which the vertical 
development and integration of the results 
can proceed and upon which a strategy for 
directed research can be developed. . 

If the principal objective addresses 
mine drainage quality then it appears 
logical to link the studies by a water 
related concept and, as a unifying theme, 
integrated by interpreting the results of 
the study within the vehicle of the 
hydrologic cycle. Consider the pathways 
of a drop of water as it enters the cycle 
as precipitation. And, depending upon the 
time of the year and the intensity and 
duration of the event, it will either run 
off and reflect the quality of the 
precipitation and surface chemistry, 
penetrate the near surface and reflect the 
quality of an actively leached horizon, or 
recharge deep within the backfill and 
reflect a quality of deep interior 
reactive spoil. Within each profile the 
variations in the degree of saturation and 
pore gas composition greatly affect the 
kinetics of alkalinity and acidity 
formation. Further, the physical 
character of the litho type, coupled with 
the method of mining, will control the 
pathway of water movement within the spoil 
(i.e., bouldery material will provide 
enhanced porosity and permeability, 
whereas clayey, friable material will 
inhibit water flow) and, in turn, the 
chemistry of the litho type will control 
the quality of the flow path. Most mine 
drainage quality predictions, based 
exclusively on overburden analyses, assume 
a uniform wetting front and ubiquitous 
water flow within the spoil, when, in 
fact, water flow is preferentially 
oriented along the more permeable zones 
(determined by the rock^s physical 
characteristics) which may be either 
acidic or alkaline. In so doing, the 



230 



majority of the water flow comprising the 
seep (and its quality) comes in contact 
with a small portion of the total backfill 
or spoil material. 

Extending this concept further to 
reclamation and in-situ acid abatement 
technology, to affect changes in chemical 
reactions, the chemistry of the water flow 
within the mine setting must be completely 
modified and controlled, an engineering 
challenge because, as discussed above, the 
runoff-infiltration ratios vary according 
to time of the year and nature of the 
precipitation. Accordingly, manipulation 
of the physical dimension is the only 
viable alternative. The number of studies 
providing for an understanding of this 
regime are remarkably few. At this stage 
of development we strive to chemically 
control a flow path by chemically altering 
the recharge point. However, within 1 
meter of penetrating the surface of a 
spoil, the direction and movement of water 
is completely unknown, confusing the 



evaluation of the efficacy of treatment. 
Did the treatment adequately intercept the 
targeted zones, was the treatment of 
sufficient magnitude, has the treatment 
gotten there yet? 

At what time will we see a call for 
innovative studies designed to stabilize 
reactive pyrites, to characterize the 
hydrology of acid producing sites leading 
toward water manipulation to control seep 
quality, to evaluate the interaction 
between revegetation, soil development and 
pores gas changes to ascertain at what 
time in the future the acidity will become 
self treating, novel ideas for the 
reconstruction of an abandoned mined land 
engineered to increase alkaline loads and 
reduce acid loads. At this level the 
unifying theme of the hydrologic cycle 
links together several elements of the 
mine drainage problem and begins to 
integrate results leading toward a central 
objective. Alternatively, do we still 
proceed to corroborate the works of 
Braley, Carpenter, Temple, Koehler, 
Starky, 



231 



STREAM SEALING TO REDUCE SURFACE WATER INFILTRATION 
INTO UNDERGROUND MINES 1 



Terry E. Ackman and J. Richard Jones2 



Abstract . --As part of the Bureau of Mines environmental 
research, a novel approach to identify and seal surface 
infiltration zones was tested at a stream, near Frostburg, MD, 
that partially overlies abandoned coal mine workings. Ground 
electromagnetic conductivity surveys were performed within a 
stream channel to identify water-saturated zones at relatively 
shallow depths of 25 and 50 ft (7.5 and 15 m). Zones of 
increased conductivity were found to be positively associated 
with areas exhibiting significant loss of flow. Conversely, 
zones which exhibited declining conductivity values delineated 
areas where there were no significant flow losses. Using this 
information, an experimental grouting procedure was used to 
place an expandable polyurethane several feet (less than a 
meter) beneath the streambed over 70 ft (23 m) of the stream 
channel. Before grouting, the section exhibited a 25 pet loss 
(800 down to 600 gal/min) ; post-grouting gaging demonstrated a 
net gain. The conductivity surveys represent a significant cost 
savings in gaging work necessary for delineating stream loss 
zones. Also, the cost of grouting was over 50 pet less than the 
costs associated with conventional rechannelization and clay- 
lining and rip-rapping techniques. 






INTRODUCTION 

Underground mining operations frequently 
induce streamflow losses, which generally 
infiltrate into underground workings and oftentimes 
become a source of perpetual water pollution (Hobba 
1981, Hollyday and McKenzie 1973, Williams et al. 
1986). Streams are considered to be the largest 
contributor to the overall volume of underground 



1 Paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the 
Interior (Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), 
April 17-22, 1988, Pittsburgh, PA. 

2 Terry E. Ackman is a Mining Engineer, and 
J. Richard Jones is a Geologist, Pittsburgh 
Research Center, Bureau of Mines, Pittsburgh, PA. 



acid mine drainage (AMD), but no quantitative work 
has ever been performed to test this assumption 
(U.S. Environmental Protection Agency 1979). In 
addition to ground water and surface water 
pollution, large water influxes into the mine can 
represent hazards to mine workers, hydrostatic 
heads can "blow out" the outcrop barrier pillars 
causing subsidence, and overflows can induce 
surface landslides. 

Partial to total streamflow losses can occur, 
and multiple infiltration zones (both natural and 
induced) usually exist in a stream channel. 
Frequently, stream losses are not readily apparent 
at the surface. It is also difficult to pinpoint 
infiltration zones from underground mines (assuming 
accessibility) since waterflow paths may deviate 
considerable distances through bedrock fracture 
systems. Although the gob or large caved areas are 
accessible through the bleeder entries, only 
limited visual observations are usually possible. 



232 



Surface rechanneling and stream channel 
liners, such as wooden canals and clay and rock 
rip-rapping, have been constructed to reduce stream 
water loss. By reducing the volume of water that 
enters underground mines from the surface, the 
volume of water discharged and treated (if 
contaminated) at the mine site should be 
considerably reduced. The cost however, of lining 
entire streams or sections of streams that overlie 
the underground workings is high. The costs for 
preventing infiltration can range between $40 and 
$80 per linear foot for relatively small streams 
(e.g., 5,000 to 10,000 gal/min (315 to 631 L/s) 
flows). Significant efforts to reduce stream 
seepage have been put forth, using the above 
mentioned techniques, in the anthracite regions of 
Pennsylvania (Ash and Whaite 1953). However, only 
limited success had been obtained by these 
techniques. The long-term effectiveness of such 
liners is questionable during times of drought or 
intermittent flow since vegetation, burrowing 
animals, and insects can affect the integrity of 
the artificial channel bottoms. The U.S. Bureau 
of Mines has, in cooperation with other Federal and 
State agencies as well as with private industry, 
instituted a research program to identify those 
sections along a stream channel that appear to be 
high subsurface infiltration zones. By pre- 
determining these high loss zones, the total length 
of stream channel that must be lined can be 
considerably reduced. Therefore, the grouting 
costs for reducing infiltration are targeted to 
those particular stream sections which have been 
identified as zones of infiltration. This paper 
reports on the methodology used to identify zones 
of subsurface infiltration, and the experimental 
results of sealing Staub Run, located near 
Frostburg, MD, with a polyurethane grout. 



ENVIRONMENTAL SETTING 

Staub Run is located in northwestern Maryland 
approximately 5 mi (8 km) south of Frostburg, MD 
(fig. 1). The section of Staub Run under study is 
approximately 0.6 mi (0.9 km) in length. Staub Run 
is considered a natural gaining stream since the 
stream does flow perennially to the point at which 
it traverses the old abandoned mined-out coal seam, 
which also outcrops in the stream channel. 
Intermittent flow is observed during the summer, 
within the stream channel section which overlies 
the abandoned workings, and drains into Georges 
Creek. Staub Run and Georges Creek are part of the 
Potomac River Watershed. Annual rainfall is about 
41 in (105 cm). Staub Run lies within the 
Appalachian Plateau Physiographic Province. The 
area is part of the Georges Creek Basin syncline, 
the northern extension of the Potomac Basin. 

A number of Pennsylvanian-aged coal-bearing 
strata underlie the area. The coal units extend 
downward from the Barton Coal within the middle 
Conemaugh Formation to the top of the Mount Savage 
Coal which marks the base of the Allegheny 
Formation (Toenges 1949, O'Hara 1900). The 
generalized thickness of these coal-bearing strata 
is some 775 ft (235 m) . 

The Pittsburgh Coal seam crops out in the 
upper portion of Staub Run. The outcrop is arcuate 
along the stream channel, such that the outcrop 
extends upslope along both sides of the mountain 
valley after surface exposure within the stream 
channel. The middle to lower portion of this small 
mountain stream overlies an abandoned turn-of-the- 
century coal mine (Carlos Mine). The depth of 
overburden for the 3,000-ft-(9!4-m-) long test site 




i% — Conductivity station 
-*• Monitoring well 

Sections chosen for grouting 



Figure 1.— Map of study area - Complete gaging 
and conductivity stations are illustrated 
on the test sections of Staub Run, 
(A-B, B-C). 



233 



ranges between and 63 ft (19 m) . Stripping 
activity (pillar recovery work) has taken place on 
both sides of Staub Run. In one portion of Staub 
Run, stripping operations have mined through the 
stream, thus necessitating re-routing of the flow 
which re-enters the natural stream channel 
approximately 1,000 ft (305 m) down gradient 
(fig. 1). The channel bottom can best be described 
as alluvial material; the sediments are sands, 
clays, pebbles, and various cobble-sized boulders. 
The thickness of this alluvial material, based on 
former and current drilling records, averages some 
12 ft (4 m) at the study site. 



METHODOLOGY 

Streamflow Gaging 

Fifteen stream gaging stations spaced at 
approximately 200-ft (60-m) intervals were 
established along Staub Run (fig. 1). Stream 
gaging began on October 23, 1986, and continued 
through September 17, 1987. Discharge was measured 
using a portable flow meter equipped with an 
electromagnetic sensor, following standard 
procedures established by the U.S. Geological 
Survey for determining the velocity/area of the 
stream. Measurements were taken at about 5-day 
intervals. 

Gaging efforts focused on establishing a flow 
profile for the study area. The initial profiling 
efforts included routine flow monitoring of gaging 
stations 1, 5, 8, 14, and 15 (fig. 1). Stations 1 
and 5 served as the control stations (outside the 
influences of mining operations) while station 15 
represented the last gaging station in the study 
area. It became apparent from gaging data, 
collected during the first full dry season, that 
significant loss zones existed in upper portions of 
the study site. This gaging data targeted the 
zone(s) for future grouting. Thus, gaging efforts 
were expanded, beginning in March 1987, to focus 
upon the upper study site by routinely monitoring 
gaging stations 1, 5, 6, 7, and 8 (fig. 1 ) , as well 
as stations 14 and 15 (fig. 2). A series of one- 
tailed t-tests were computed to test if the 
downstream station discharge was significantly 
greater (p < 0.05) than the corresponding upstream 
station discharge. 



1. 8 
1.7 
1.6 
1.5 



I 1-3 



1.2 

I.I 



100 200 300 400 500 600 700 800 900 
DISTANCE FROM STATION 1, m 

Figure 2. --Mean annual discharge at Staub Run. 



1 


1 1 f 


i i I 


I l 


- 


- 








15 


. 






14 


— i' 


- 


>v5 






- 


- 


KEY * 

• Gaging station 






- 




i i i 


. i i 


i ' 





Electromagnetic Ground Conductivity Surveys 

Electromagnetic ground conductivity surveys 
were used in a novel attempt to identify stream 
loss zones by delineating the presence of water- 
saturated zones directly beneath the stream 
channel. The equipment consisted of a 
transmitter, transmitter coil (wire loop), 
receiver, receiver coil, and connecting cable. A 
two-person crew is required for using this entirely 
portable equipment. One person carries the 
transmitter and transmitter coil and the other 
person carries the receiver and receiver coil. 
Readings were taken at a fixed distance of 33 ft 
(10 m) between transmitter and receiver coils; 
however, fixed distances of 66 ft (20 m) and 132 ft 
(40 m) are also available for greater depths of 
penetration with this equipment. 

The surface transmitter induces a current in 
the subsurface material with this electromagnetic 
induction technique. An alternating magnetic field 
is produced from the alternating current generated 
from the transmitter coil. The magnetic field 
induces current flow through the substrate as it 
penetrates the ground surface. The receiver coil 
senses a secondary magnetic field which is 
generated by the induced currents. The secondary 
magnetic field sensed at the coil is a function of: 
the strength of the primary field; current 
frequency, distance between transmitting and 
receiving coils, and ground conductivity. The 
ground conductivity is the only unknown variable 
since the primary field, frequency, and coil 
separation can be controlled. The receiver, which 
senses the secondary magnetic field, internally 
converts the signal to terrain conductivity and 
digitally displays signal in millimhos/meter 
(mmhos/m) . The reader is referred to standard 
geophysical texts for detailed derivations (McNeill 
1980, Grant and West 1965, Keller and Frischknecht 
1966, Ladwig 1982) since the mathematical theory 
behind the induction technique is beyond the scope 
of this paper. 

Theoretically, the magnitude of the ground 
conductivity should increase when saturated 
conditions exist. Consequently, the water loss 
zones should be zones of high conductivity. Thus, 
the conductivity data and gaging data should have 
an inverse relationship. 

Two series of conductivity measurements were 
taken within the stream channel between the 15 
gaging stations at 33 _ ft (10-m) spacings. One 
series of measurements was taken during the wet 
season and the other during the dry season (figs. 3 
and 4). A Geonics EM-3 1 * electromagnetic ground 
conductivity meter was used to obtain measurements 
at about 25- and 50-ft (7.5- and 15-m) depths. The 
observation depth is changed by changing the 
orientation of the instrument. Several portions of 
Staub Run, between conductivity stations 2 and 5 
and between stations 64 and 69, may have registered 
erroneous readings with the conductivity instrument 
due to metallic interference (steel pipes, old 
appliances, and metal debris). 

A statistical comparison between the 
conductivity and gaging data was performed. The 
conductivity data from the wet and dry seasons were 
combined to generate an annual mean conductivity 
(fig. 5). This was done by taking the difference 



234 



10 



8 4 

Q 

z 
o 
o 



Dry season 
Gaging station 5 




KEY 
• Measured at 7.5 m depth 
A Measured at 1 5 m depth 



100 200 300 400 500 600 700 800 900 
DISTANCE FROM STATION 1, m 



Figure 3. --Conductivity survey (dry season) at 
Staub Run. 



o 

Q 
O 



\ot 

8 
6 
4h 



1 

Wet 


I 1 I 1 1— 

season 

/-Gaging station 5 


i 


i 


7 




69 




- \ 


A * 6 7 Conductivity 
/ AA. station 64 — • 

/ ^\t. 8 A 


l\ 


f A. *- 


t 


vA \tj 


Xii 


fa- 


\ 


J V^r r 


..; 


ir- 


^ 


• Measured at 75m depth 


Y 




i 


A Measured at 15m depth 

i i i i 1— 


i 


■ 



100 200 300 400 500 600 700 800 900 
DISTANCE FROM STATION 1, m 

Figure 4. --Conductivity survey (wet season) at 
Staub Run. 

between the two observation depths as a means of 
normalizing the data. The differences between the 
vertical and horizontal readings for each station 
of both series were then added together and divided 
by 2 to obtain a mean conductivity. 



Corresponding measurements were most complete 
between gaging stations 1 through 8 and 
conductivity stations 1 through 45, so these data 
were used to generate the respective polynomial 
equations. Furthermore, the mean station gaging 
discharge (October 23, 1986 to September 17, 1987) 
and the mean station conductivity measurement (wet 
season, dry season) are considered representative 
of the seasonal trends so the respective averages 
are used as the deDendent variables with the 
conductivity station locations serving as the 
independent variables. After determination of the 
best-fit equation for the respective data, 
predictive values are generated for the station 
measurements and then compared with Spearman's Rank 
Order correlation analysis. 



Grouting Procedure 

Experimental grouting was conducted between 
gaging stations 11 and 15 (fig. 1). Grout rods 
3 ft (1 m) in length were used for injection of the 
polyurethane material into the alluvial stream 
sediments. The rods were 3/4-in (1.9-cm) diameter 
steel pipes with hardened steel points with 1/8-in 
(0.3-cm) holes drilled through the sidewalls near 
the lower portion of the rod. The top of the 
threaded rods incorporated a hardened steel cap 
which would withstand blows from a sledge hammer 
and could also be used for mechanical injection of 
the polyurethane grout. The rods were manually 
driven into the stream sediments with a sledge 
hammer to a uniform depth of 2 ft (0.66 m) beneath 
the stream channel and placed at 10-ft (3-m) 
centers. 

A measure of 5 gal (19 L) or 44.5 lbs 
(20.2 kg) of a two-component polyurethane grout, 
2.5 gal (9.5 L) of each component, was injected 
into each grout rod. Injection pressures ranged 
from 400 to 1,200 lb/in 2 (28 to 84 kg/cm 2 ) with the 
in situ conditions of alluvial material dictating 
the pressure. The pump which was used was capable 
of 2,000 lb/in 2 psi (141 kg/cm 2 ). Grout was 
injected at a maximum pressure obtainable until a 
surface leak developed. After the leak occurred, 
the injection pressure was reduced to 50 lb/in 2 
(3 kg/cm 2 ) until the leak sealed itself. The 
pressure was then increased slowly until another 
leak occurred or the 5 gal (19 L) of material was 
depleted. 



WET DRY 
Mean Conductivity = ((V 



H) + (V - H))/2 



Since the number of stations where conductivity 
measurements were recorded exceeded those where 
stream gaging occurred (figs. 1 and 2), comparison 
between the two data sets was accomplished by using 
polynomials to predict the statistically best-fit 
equation for the respective data. In this way the 
conductivity measurements and the gaging records 
were each treated as separate mathematical curves. 
Once a polynomial curve is fit to the respective 
data, a predicted trend pattern emerges between 
the two measures whose station to station 
relationships can be further tested. The 
mathematics and examples of polynomial curve 
fitting applications in the earth sciences are 
numerous (Doornkamp and King 1971 , Jones and 
Cameron 1977, and Fisher et al (in press)). 



RESULTS 

Gaging and Conductivity 

In most natural streams there should be a 
downstream increase in the volume of discharge. To 
test if the between- station volumes along 
Staub Run followed this natural pattern, a series 
of one-tailed t-tests was computed for the 
contiguous gaging stations 1 to 5, 5 to 6, 6 to 7, 
7 to 8, and 5 to 8 (table 1). The one-tailed 
t-test not only determines if a significant 
difference (p <_ 0.05) exists in discharge 
measurements between the respective gaging 
stations, but is directional in that one can test 
if the discharge at one gaging station is 
significantly greater than that of another gaging 
station. 



235 



4 - 



> 

I- 3 
o 

Q 

Z 

o ? 



1 1 1 T" 

KEY 


i i 


ft 




• Annual mean 




i 




- 




M 




A 6 A 




y 




rl/|jv \ £r 




m 




■ / * \ v^ 




In 




f\ 4 Gaging station 5 


T9f J 


n 


J 


■ .i 1 1 


_i l 


|I4" 


T 

15 



100 200 300 400 500 600 700 800 900 
DISTANCE FROM STATION 1, m 

Figure 5. --Mean annual conductivity Staub Run. 

Since flow rates should increase in a 
downstream direction, the test was formulated such 
that the contiguous downstream station discharge 
should be significantly greater than the 
corresponding upstream station discharge. If it 
can be demonstrated that the reverse is 
statistically significant, i.e., the upstream 
station has larger discharges, then it is assumed 
that the stream segment between the two gaging 
stations is experiencing a flow loss and thus 
represents a zone of significant subsurface 
infiltration. For the t-tests between the two 
contiguous stations, discharge data are only used 
when measurements were recorded on the same day at 
the two stations (table 1). 

The results of the t-tests are interpreted to 
show that significant downstream losses occur 
between stations 1 and 5, and stations 6 and 7. 
Significantly greater upstream discharges were not 
identified between stations 5 and 6 and stations 7 
and 8 (table 1). Following the original 
hypothesis, the stream segments located between 
stations 1 and 5 and stations 6 and 7 represent 
zones of statistically significant subsurface 
infiltration. Although the downstream flow rates 
are apparently less than the corresponding upstream 
flow rates between stations 5 and 6 and stations 7 
and 8 (table 1), the differences are not 
statistically significant (p <_0.05). To further 
test an overall flow loss pattern, a one-tailed t- 
test was also computed between stations 5 and 8, 
again using only those gaging measurements when 
recorded on the same day at the respective 



stations (table 1). The t-test results illustrate 
a significantly (p <_0.05) higher flow at station 5 
as compared to the flow at station 8 (fig. 2). 
This appears to confirm that infiltration zones are 
occurring downstream between station 5 and 
station 8 along this portion of Staub Run. 

As noted in table 1 , there is a significant 
discharge loss occurring between gaging stations 1 
and 5 (fig. 2). The local bedrock and structural 
geology indicate that coal could not have been 
extracted from beneath this area. Examination of 
stereo paired aerial photographs (1:1000, 1983 
series) appears to show the presence of a linear, 
crossing the stream in the vicinity of gaging 
station 5. The area of the apparent linear has, 
however, been significantly altered by man through 
surface mining, timbering, and the construction of 
residential and other types of structures so it is 
difficult to confirm its existence. Field 
reconnaissance of the area was also unable to 
follow a continuous linear trace. 

The preliminary analysis of the gaging data 
and conductivity surveys does show a significant 
loss zone in this area so additional analysis of 
the aerial photographs for confirmation of the 
linear is presently being made. 

The discharge measurements are best defined by 
a second-degree equation. The multiple coefficient 
of correlation (R 2 ) for the second-degree equation 
is 0.94. The predicted discharge values by station 
are presented in table 2 with the predicted curve 
trend illustrated in figure 6. 

The mean conductivity measures are best 
defined by a sixth-degree equation. The multiple 
coefficient of correlation (R 2 ) for the sixth- 
degree equation is 0.71. The predicted 
conductivity measurements by station are presented 
in table 2 with the predicted curve trend 
illustrated in figure 6. 

Comparison of the predicted curve trends 
between the discharge and conductivity measurements 
appear to show a negative relationship (fig. 6), 
i.e., a station increase in conductivity appears to 
correspond to a decrease in station discharge. To 
test if this apparent trend is significant, the 
non-parametric Spearman's Rank Order Correlation 
was computed (Doornkamp and King 1971). 






Table 1. --Paired t-tests for gaging data. 



Statistical 

pairs (gaging 

stations ) 


Average 
flow, 
gal/m 


Maximum 
flow, 
gal/m 


Minimum 
flow, 
gal/m 


Sum of 

flows , 

gal/m 


Number of 
gaging 
events 


Statistically 
significant 


1 
5 


1,060 
92 3 


4,353 
3,62 9 


6 
5 


45,589 
39,687 


43 


Yes 


5 
6 


142*1 
1409 


1,695 
1,384 


5 
8 


5,939 
5,720 


14 


No 


6 

7 


394 
32 7 


327 
1,114 


35 

21 


5,905 
4,90 5 


15 


Yes 


7 
8 


1471 
14 62 


1,114 
1,249 


35 

1 


4,713 
4,616 


10 


No 


5 
8 


94 6 
696 


3,629 
2,50 6 


64 

1 


34,986 
25,763 


37 


Yes 



236 



' 1 ' 1 


i | 1 | 1 


r* 




AiAAii 

AAA1 

A AA 

— AA 


.?1""-. . 


AAA 


• • 


A 


• 




»» • • 


_ 


".: 




AA • 




• A 




• AA 




• A.. • 


• 
• 


I 1 *-. 


• 


A 


• 


A 


• • 


*A 


• 


A 


• • 




• •• 




— ••• 






KEY 




a Discharge value 


i . I 


• Conductivity value 
i 1 i 1 . 


10 20 


30 40 50 


CONDUCTIVITY STATIONS „».„ 



Figure 6. --Predicted curve trends between the 
discharge and conductivity measurements. The 
predicted second-degree equation for the 
discharge is given by: y = 438.16 - 0.887x - 
0.084x 2 , where y = predicted mean discharge and 
x = station location. The predicted sixth- 
degree equation for the conductivity is given 
by: v = 1.85 + 0.013x + 11.97x 2 + 0.00008x 3 - 
17. 7x* - 18.00x 5 - 0.0000000009X 5 , where 
y = predicted mean conductivity measurement and 
x = station location. 



The Spearman's Rank Order Correlation 
coefficient is -0.92 suggesting a strongly negative 
relationship between the station discharge and 
conductivity measurements. Thus, the predicted 
curve trends of the two measurements illustrated in 
figure 6 do show a negative relationship, such that 
high conductivity readings generally correspond to 
lower discharge measurements. In terms of 
identifying subsurface infiltration zones, the 
stations exhibiting lower discharges and higher 
conductivity measurements are areas where stream- 
flows are infiltrating into the underground 
workings. Alternatively, the stations exhibiting 
higher discharges and lower conductivity 
measurements are areas where stream water does not 
appear to be infiltrating or there is less 
infiltration into the mine. 



Grouting 

In order to establish a methodology for 
surface grout injection, a section of stream 
channel between conductivity stations 82 and 84 was 
selected as the test area (fig. 1). Prior to the 
grout injection procedure previously described 
discharge measurements were taken at conductivity 
stations 82 and 84 on May 20 and May 28, 1987 
(table 3). The polyurethane grout was injected 
into the stream at 10-ft (3-3-m) centers between 
conductivity stations 82 and 84, i.e., seven grout 
rods were used along the 70-ft (23-m) section of 
stream following gaging on May 28, 1987. Post- 
grouting gaging was conducted on May 29, June 4, 
and June 5, 1987 (table 3). As illustrated in 
table 3, the discharge rates between conductivity 
stations 82 and 84 showed a net gain in flow after 
grouting was completed. 

Unfortunately, stream discharge decreased to 
low flow conditions within the grout test area 
around mid-June 1987 (approach of the dry season) 
and eventually the stream flow terminated. The 



overall gain in discharge between these two stream 
points for the time of gaging does, however, 
suggest that the grout injection procedure does 
prevent subsurface infiltration. 

Based on this limited sealing operation, costs 
are estimated to range between $25 and $30 per 
linear foot of the 10-ft (3-3-m) wide stream 
channel. Other streams in the watershed of 
approximately the same size and geological 
conditions have been rechannelized with heavy 
equipment and sealed with clay-lining and rock rip- 
rapping. The costs of these sealing operations 
were over $70 per linear foot. In addition to 
significantly disturbing natural conditions, the 
effectiveness of the latter technique is 
questionable based on visual observations of the 
streams during various flow conditions. Future 
evaluation of the conventional stream sealing 
techniques are scheduled. 



CONCLUSIONS 

This novel approach to identifying 
infiltration zones has demonstrated that areas of 
high conductivity trends were associated with 
significant stream losses, and that areas of 
declining conductivity trends demonstrated no 
significant losses in stream flows. This technique 
demonstrates a potential for accurately locating 
stream loss zones without flow monitoring for a 
full hydrological year. The two conductivity 
surveys (wet and dry season) over the 3,000-ft 
(910-m) stream segment took only a day each to 
complete. Thus, by performing conductivity 
surveys coupled with confirmation gaging, areas of 
stream infiltration were located with minimal time 
and effort. In addition, this pinpointing of loss 
zones was indiscriminate of natural or manmade 
causes. This implies that the potential for 
identifying natural fracture zones (linears) 
exists, and this technique may be a predictive tool 
for possible roof control problems in active 
underground workings. 

The stream sealing technique which has been 
developed was demonstrated to be quick and easy, 
causing minimal disturbance to the natural 
conditions." This novel approach shows good 
potential for reducing surface and ground water 
infiltration into underground workings (both active 
and abandoned). Limited data have been encouraging 
in terms of effectiveness. In addition, the 
potential for 50 pet cost savings exists when 
compared to conventional clay-lining and rock rip- 
rapping techniques. Consequently, this technique 
offers an economical means of reducing pollution in 
abandoned mines, as well as reducing water-handling 
and treatment costs in active mines. 

Based on data collected thus far, full-scale 
grouting targets have been identified in the upper 
project site between conductivity stations 20 to 27 
and 32 to 45 (fig.1). This grouting work is 
scheduled to be completed in the near future and 
will affect 800 ft (242 m) of stream channel while 
only actually sealing 600 ft (182 m) . Although 
this project appears promising for alluvial type 
stream channels, more work needs to be performed 
with bedrock stream channels. In addition, the 
bedrock situation is anticipated to be slightly 
more expensive due to obvious drilling operations 
which would be required. 



237 



Table 2. 


-- Predicted mean conductivity and mean discharge 




measurements . 










Predicted 


Predicted 


Mean Measured 


Conductivity 


Gaging 


conductivity, 


di scharge , 


discharge 


station 


station 


mmho/m 


gal/m 


gal/m 


1 


1 


1.86 


437.19 


433 


2 




1.88 


436.05 




3 




1.89 


434.74 




4 




1.91 


433-27 




5 




1.92 


431.63 




6 




1.95 


429-81 




7 




1.97 


427.83 




8 




1.99 


425.69 




9 




2.02 


423-37 




10 




2.06 


420.89 




11 




2.10 


418.24 




12 




2. 14 


415.42 




13 




2.1 9 


41 2.43 




14 




2.25 


409.28 




15 




2.31 


405.96 




16 




2.37 


402.46 




17 




2.44 


398.81 




18 




2.52 


394 . 98 




19 




2.61 


390.98 




20 




2.70 


386 . 82 




21 




2.80 


382.49 




22 




2.90 


377.99 




23 




3.00 


373-32 




24 




3. 12 


368.49 




25 




3-23 


363.49 




26 


5 


3-35 


358.31 


394 


27 




3.47 


352.97 




28 




3.58 


347.47 




29 




3.70 


341.79 




30 




3.82 


335.95 




31 




3.93 


329.94 




32 


6 


4.03 


323.76 


305 


33 




4.1 2 


317.41 




34 




4.20 


310.90 




35 




4.26 


304.21 




36 




4.31 


297-36 




37 




4.33 


290.35 




38 


7 


4.33 


283 . 1 6 


250 


39 




4.29 


275.80 




40 




4.21 


268.28 




41 




4.10 


260.59 




42 




3.93 


252 . 73 




43 




3-71 


244.70 




44 




3.43 


236.51 




45 


8 


3.08 


228.1 4 


249 



Table 3. 



Pregrouting and post-grouting discharge change. 



May 20, 1 987 

May 28, 1987 

*May 2 9, 1987 

*June 4, 1987 

*June 5, 1 987 



Conductivity 
station 82 
discharge , 
gal/m 



Conductivity 

station 84 

discharge, 

gal/m 



C nan ge , 
gal/m 



1,228 
845 
618 
237 
262 



1,194 
634 
693 
332 
280 



-34 
-211 
+75 
+95 
+18 



*P os t- grouting discharges. 



238 



LITERATURE CITED 

Ash, S.H. and R. H. Whaite, 1953- Surface-Water 

seepage into anthracite Mines into the Wyoming 
Basin Northern Field. BuMines Bulletin 534, 
30 pp. 

Doornkamp, J. C. and King, C. A., M. , 1971. 
Numerical analysis in geomorphology. 
St. Martins Press, New York, NY, 353 pp. 

Fisher, J. J., J. R. Jones, and E. Tynan. Sediment 
distribution along the Rhode Island southshore 
beaches. Physical Geography . In Press. 

Grant, F. S. and G. F. West, 1965. Interpretation 
theory in applied geophysics. McGraw-Hill 
Book Co., New York, 583 pp. 

Hobba, W. A., 1981. Effects of underground mining 
and mine collapse on hydrology of selected 
basins in West Virginia. U.S. Geological 
Survey RI-33, 72 pp. 

Hollyday, E. F. and S. W. McKenzie, 1973. 
Hydrogeology of the formation and 
neutralization of acid waters draining from 
underground coal mines of western Maryland. 
Maryland Geological Survey, RI-No. 20, 50 pp. 

Jones, J. R. and B. Cameron, 1977. Landward 

migration of barrier island sands under stable 
sea-level conditions: Plum Island, 
Massachusetts. J. of Sedimentary Petrology, 
vol. 47, pp. 1475-1483. 

Keller, G. V. and F. C. Frischknecht, 1966. 

Electrical methods in geophysical prospecting. 
Permagon Press, New York, 519 pp. 

Ladwig, K. J, 1982. Delineation of zones of acid 
mine drainage using surface geophysics. In 
Proceedings, Symposium on Surface Mining, 
Hydrology, Sedimentology, and Reclamation, 
Univ. of KY, pp. 279-287. 

McNeill, J. D, 1980. Electromagnetic terrain 
conductivity measurement at low induction 
numbers. Technical Note 6, Geonics Limited, 
1745 Meyerside Drive, Mississauga, Ontario, 
Canada. 

O'Hara, C. C, 1900. The geology of Allegheny 

County. Maryland Geological Survey, 159 pp. 

Toenges, Albert L., 1949. Investigation of lower 
coal beds in Georges Creek and northern part 
of upper Potomac Basins, Allegheny and Garrett 
Counties, MD. BuMines Technical Paper 725, 
100 pp. 

U.S. Environmental Protection Agency, 1979. 

Dewatering active underground coal mines: 
technical aspects and cost-effectiveness. 
EPA-600/7-79-124, 123 pp. 

Williams, R. E., G. V. Winter, G. L. Bloomsburg, 
and D. R. Ralston, 1986. Mine hydrology. 
Society of Mining Engineers Inc., 169 pp. 



239 



UNSATURATED FLUID FLOW IN MINE SPOIL: 
INVESTIGATIVE METHODS LEADING TO A QUANTITATIVE CHARACTERIZATION 



David M. Diodato and Richard R. Parizek 



A tracer study using non-radioactive neutron activatable bromide-79 was 
employed as an investigative technique to determine in situ unsaturated 
hydraulic conductivities at a reclaimed and revegetated strip mine in Clarion, 
PA. This investigation is a part of an ongoing acid mine drainage abatement 
study using surficial lime plant flue dust and limestone quarry waste. Water 
samples were collected through time from 55 pressure-suction lysimeters 
installed at various depths at the field area. 23 were selected for study and 
neutron activation analysis was used to determine Br concentration. 
Interpretation of the concentration peaks as average arrival times of infiltrating 
rainwater yielded unsaturated hydraulic conductivities ranging from 0.0279 to 
0.5313 ft/day with one value of 1.1625 ft/day. This high value is interpreted as 
an example of infiltrating water piping through the highly heterogeneous 
disturbed spoil pile. Excluding the high and low values, the mean and the 
median unsaturated hydraulic conductivities were found to be 0.1876 ft/day and 
0.1582 ft/day, respectively. These values are in the range of 0.14 to 18.76 
percent of saturated hydraulic conductivity of the various geologic materials at 
the site. Further refinement of the data may be yielded by geostatistical 
investigations, as the semivariogram of the concentration values at a sampling 
point through time may suggest alternative concentration peak dates. A second 
investigative technique, the use of neutron depth-density and depth-moisture 
probes, will be of use in the determination of soil bulk density and soil moisture 
content values. An increased understanding of the behavior of alkaline waters 
and the hydraulic conductivity in the unsaturated zone will help investigators to 
evaluate past abatement strategies and to plan future ones. 



INTRODUCTION 

The unsaturated hydraulic conductivities of a reclaimed 
and revegetated acid-producing strip mine are being investigated 
and quantified by means of a tracer experiment. This research is 
in conjunction with a large-scale acid mine drainage abatement 
experiment the C&K Coal Old Forty Strip Mine site, Clarion, 
PA. The latter seeks to quantify the effects of two surficial 
treatments, each using different application amounts of lime 
plant flue dust and limestone quarry waste, in abating the acid 
production of the strip mine. On a macroscopic level, the 
alkaline waters resulting from precipitation on the applications in 
the two treatment plots may have been expected to percolate 
down through the spoil pile as a relatively uniform front. 
However, inspection of chemical analyses of soil water samples 
from the site suggested that the downward flow of these waters 
was not uniform. A desire for an increased understanding of the 
nature of unsaturated fluid flow at the site was the motivation 
for the bromide tracer experiment. 



David M. Diodato is a Graduate Research Assistant and 
Richard R. Parizek is a Professor of Geology, Dept of 
Geosciences, The Pennsylvania State University, University 
Park, PA. 



20 40 60 80 100 120 




Border Values in 100s of FT 



Figure 1. Site map showing limits of mining, treatment plots 
1 and 2. and the control plot, C. 



240 



At the site there are 17 drill holes that contain nests of 
pressure-suction lysimeters. These are buried at approximately 
7-ft depth intervals. Figure one is a map of the site showing the 
treatment plots and lysimeter nest locations. The drill holes 
were backfilled preferentially using drill cuttings and other spoil 
materials readily available at the site. In some cases, it was 
necessary to use additional, exotic sand and gravel. But, in 
general, the use of materials found at the site helped to ensure 
that drill holes remained representative, both chemically and 
hydraulically, of the adjacent spoil pile. 




a) 



Lysimeter 
Nest 



SampL 
Tube 




Pressure/Suction 
Tube 
Plastic 

Cylinder/ 



Ceramic 



Cup 



b) 



Figure 2. a) Schematic pressure-suction lysimeter nest 

installation, b) Cut-away view of a pressure-suction 
lvsimeter. 



Figure two is a schematic of a lysimeter nest installation. Soil 
water samples have been collected from these devices on a bi- 
weekly to monthly basis since 1984. What has been observed in 
some of these nests is an increased alkalinity at shallow depths, 
more acidic conditions at intermediate depths, and increased 
alkalinity at greater depths. This anomalous result immediately 
suggests three hypotheses. One hypothesis is that these lower 
alkalinity and pH samples could be produced by isolated hot 
spots in the spoil pile. If this is the case, then optimal remedial 
efforts should center on the identification and treatment of these 
hot spots, which may be responsible for the bulk of acid 
production. Alternatively, these low alkalinity zones could reflect 
decreased circulation of alkaline charged waters resulting from 
the treatment. This might result if these zones were more shaly, 
the disaggregated shale having a lower hydraulic conductivity 
than adjacent disaggregated blocky sandstone. Finally, this 
finding could be a result of a combination of the two previously 
mentioned factors, with the shaly zones containing the bulk of 
the pyritic material. The second hypothesis may be evaluated by 
determining the in situ unsaturated hydraulic conductivity of 
many zones in the spoil pile. In addition, knowledge of 
representative unsaturated hydraulic conductivity values will be 
of use to workers in this field. 



THEORETICAL BACKGROUND 

A tracer experiment is underway which aims to map 
hydraulic conductivities through the spoil pile. The chosen tracer 
was the non-radioactive, neutron activatable bromide-79. The 
advantages of bromide-79 as a tracer are outlined in the 
literature Schmotzer et al. (1973) and Jester et al. (1977). 
Briefly, they are as follows. Background bromide typically exists 
in natural aqueous systems at very low levels (40 to 300 ppb) in 
geologic association with evaporitic brine and marine shale 
sequences. Neutron activation analyses of background samples 
from the site showed no detectable bromide. It is not toxic to 
humans below 0.5 to 1.0 mg/mL of blood. Bromide is 
conservative in that it does not readily adsorb, absorb, or 
precipitate to a significant degree on natural surfaces. Bromide 



is unaffected by the actions of microbial biota. Further, the 
sensitivity of detectors allows resolution of background bromide 
to as low as 20 ppb. In addition to this, analysis and initial 
purchase cost of bromide are relatively inexpensive (Schmotzer et 
al. 1973). 25 pounds of potassium bromide were purchased for 
this experiment for less than two hundred dollars. Ammonia 
bromide is even less expensive, but potassium bromide was 
chosen to eliminate any slight nitrification of ground waters 
which may otherwise have occurred. 

The quantification of unsaturated hydraulic conductivity in 
this study is accomplished by direct measurement of in situ 
bromide contents of infiltrating waters through time, the peaks 
of which are inferred to represent average transport times. This 
inference is not without precedent (Schmotzer et al. 1973, Jester 
et al. 1977, Brasino and Hoopes 1985, Raupach et al. 1983). 

Saturated Hydraulic Conductivity 

Hydraulic conductivity is a parameter of the D'Arcy fluid 
flow relation which describes the capacity of a given medium to 
transmit a given fluid. It is expressed in the units of a velocity, 
length per time. Hydraulic conductivity is a function of the 
characteristics of both the medium and the fluid moving through 
it. This is well illustrated by the relation 



K = ( K p g )/ |X 



(1) 



where 



K = Hydraulic Conductivity 
k = Intrinsic Permeability 
p = Fluid Density 
g = Acceleration Due to Gravity 
p = Fluid Dynamic Viscosity 

(Freeze and Cherry 1979) 

Here the intrinsic permeability, k , is a function of the medium 
alone, and includes factors describing mean grain diameter, the 
packing arrangement of grains, and the sphericity and roundness 
of grains, among other factors. It is a term best suited to the 
idealized D'Arcian elemental volume concept, and less well suited 
to more complex fluid flow such as that which exists in fracture 
flow controlled terrains. The terms which describe fluid 
properties are density, p , and dynamic viscosity, p.. These 
properties are well known and tabulated for a number of fluids 
at standard temperature and pressure, and for water at 
standard pressure over a wide range of temperatures. For water 
flow in a near-surface, non-hydrothermal, non-brine 
environment, these properties may be safely assumed to be 
constant. Similarly, the earth's gravitational force, which is 
acting on the fluid, decreases exponentially with distance away 
from the core, but may be considered constant within a small 
range of elevations. 

Unsaturated Hydraulic Conductivity 

For any given medium and fluid, saturated hydraulic 
conductivity represents the limiting case in that it is the 
maximum measure of this fluid transmitting capacity. As a 
medium becomes increasingly unsaturated with respect to a 
fluid, the volume of the medium transmitting the fluid decreases. 
Because there is a loss of fluid transmitting volume measured in 
L 3 , expressed in a term of L/T, the decrease in hydraulic 
conductivity with increasing unsaturation is often dramatic and 
may span several orders of magnitude or more. 

The degree of unsaturation can be expressed by a number 
of parameters, two of which are volumetric water content, 9 , 
and matric potential, V P. Volumetric water content is the volume 
of water in a unit volume of medium, commonly expressed as a 
percentage. Thus if a medium has a porosity of 20 percent, and 
a volumetric water content of 20 percent, it is fully saturated. 
Increasing air content reflects increasing unsaturation. It should 
be noted that porosity does not imply permeability, because the 
void spaces are not necessarily interconnected. 



241 



Alternatively, matric potential is a measure of the tension 
at which water is held within the medium. In the literature, the 
symbol H 7 is commonly used to express the pressure head portion 
of the total hydraulic head relation, and thus includes both 
positive and negative pressure regimes. Here, we use v r" to 
indicate only the nonpositive pressures which exist in the 
unsaturated zone. In adherence with established conventions, 
matric potential pressures are expressed as equivalent height of 
a column of water. This is a function of the mean grain size of 
the medium, among a number of other factors, and can be 
expressed by a capillarity relation. 



where 



* = ( 2 o cos9 )/( r p g ) 



(2) 



T = Height of Capillary Rise 

o - The Surface Tension of the Water-Rock Interface 

9 = The Angle of Contact of the Capillary Force 

r = The Radius of the Pore 

p = Fluid Density 

g = Acceleration Due to Gravity (after Miller 1982) 

Consistent with potential theory, fluid will flow from higher to 
lower matric potential pressures. Thus one way to view matric 
potential is as the nonpositive part of the continuum between 
negative and positive fluid pressures, with zero matric potential 
at the water table. Commonly, above the water table there 
exists a tension-saturated zone where water is stored in the 
medium under capillary force. Both matric potential and 
volumetric water content can be measured with field 
instrumentation and are of direct bearing on unsaturated 
hydraulic conductivity. Thus, unsaturated hydraulic conductivity 
is commonly expressed as a function of matric potential, K(& 
V P& ), or as a function of volumetric water content, K(& 6). By 
either measure, saturated hydraulic conductivity is the upward 
limiting case. 

The phenomenon of unsaturation has many important and 
complicating consequences. Among these, one of theoretical 
significance is the behavior of a slug of water associated with a 
single infiltration event. This slug will have a leading wetting 
front and a trailing drying front as it percolates downward 
through the unsaturated zone. A graph of either volumetric 
moisture content or matric potential versus unsaturated 
hydraulic conductivity has two hysteretic curves, one describing 
each front. This is because the wetting front has a greater 
hydraulic conductivity than the drying front. Because the tracer 
experiment is not strictly infiltration event-oriented, an 
investigation into this transient phenomenon is beyond the scope 
of this study. 



EXPERIMENTAL METHODS 

Two sets of background samples were collected from the 
55 pressure-suction lysimeters installed at the site, the first on 
November 5, 1985 and the second on April 5, 1986. 
Subsequently, on April 4, 1986, a 4,080 mg/L aqueous bromide 
solution was applied around each of the lysimeter nests, except 
for TR2LY5, which has only one functioning sampling device. 
The application pattern is illustrated schematically in figure 3. 
The application was in a radial hexaseptate pattern designed to 
emulate a diffuse source. The bromide application emulated a 
diffuse source in an effort to mimic the behavior of both the 
rainwater and the limestone quarry waste and flue dust 
abatement application, all of which are diffuse sources. Each 
radius was 25 ft in length and received 5 gal of solution. Care 
was taken to avoid application immediately on the lysimeter 
nest, so as to avoid any possibility of downhole piping or 
channeling through backfill sediments. Soil and meteorological 
conditions at the time of the application were ideal. The land 
surface was dry and rapid infiltration occurred with no runoff 
observed. Coincident with the completion of the application, a 
very gentle rain began to fall. This drizzle increased in intensity 
at a slow and steady rate over the next 6 to 12 hours, ensuring 
maximum infiltration of the tracer. 



1300 




25 feet 
Lysimeter Neat 



Radial Hexaseptate 
Application Pattern 



Figure 3. Schematic of tracer application pattern with 
contour lines. 



Water samples were collected from the pressure-vacuum 
lysimeters installed in the 17 lysimeter nests at the site, from 
surface waters which flow through and off the site, and from 
ground water wells. The collection schedule varied according to 
precipitation and evapotranspiration, but averaged about three 
times a month during the initial four month period beginning 
April 14, 1986. For the next eight months sampling was 
conducted on a monthly basis. After April 14, 1987, one full 
year of bromide sampling had been completed, and the sampling 
schedule was reduced to a bimonthly one. The last sampling was 
in August 1987. A careful sampling procedure ensured that the 
integrity of the samples is maintained during collection. Because 
of the stability of the bromide anion, samples are safely stored at 
room temperature for several months. 

The sampling procedure involves several anti- 
contamination safeguards. Sample bottles are kept sealed prior 
to and following sampling. While the lysimeter sampling tubes 
are held earthward, they are rinsed with distilled water and 
dried with paper towels. This cleaning is designed to eliminate 
the presence of contaminants which may have been derived from 
surficial sources. Initially, as the sample is being drawn, a small 
volume is discharged to the ground in order to further flush the 
sample tube. During sampling, the lids of the bottles are kept off 
the ground, as the ground potentially contains very high levels of 
Br". The sample is collected in the bottle, taking care to prevent 
the outside of the sample tube from coming in contact with the 
inside of the bottle, and the lid of the bottle is reattatched. 

Neutron activation analysis is a straightforward 
procedure. Five mL portions of the samples are placed in 2 dr 
poly vials and irradiated with gamma radiation. Neutron 
activation produces a number of short-lived isotopes of 
bromide-79, including bromide-80 and bromide-82. A 50 cm 3 
ge(li) detector counts the disintegrations, and feeds the data to an 
ND-680 pulse height analyzer. Because each isotope has a 
unique spectrum of escape energies, isotopes may be readily 
identified. By comparing the magnitude of peaks of a particular 
escape energy with the magnitude of peaks of the same escape 
energy produced by a standard of known concentration, 
concentrations can easily be calculated. 

The method of calculation for bromide concentration in 
neutron activated water is simplified by holding several of the 
variables constant. For any given run, all samples, including 
standards, are activated at a constant flux energy level for the 
same amount of time, and counted with the detector mounted at 
a constant distance for the same amount of time. A correction 
must be made for decay times for all of the samples, including 
the standards. Most sample group activations were performed 
at 900 kW for 1 minute, although some were at 1,000 kW for 1 
minute. In every case, counts lasted 45 minutes. The governing 



242 



relation for radioactive decay is 

N = N exp[-Xt/T 1/2 ] (4) 

where 

N = The observed sample activity after decay time, t. 
N = The initial sample activity. 
X = A decay constant, 0.693. 
t = Decay time, time elapsed between sample 
activation and counting. 
T1/2 = The half-life of the isotope of interest. 

Having completed a series of activations and counts, the initial 
activity of the standard must be calculated. Rearranging (4), 

N = N exp[Xt/T 1/2 ] (5) 

Here, N is the number of counts that the 50.0 parts per million 
standard would have produced at t = 0. Next, the initial 
activity of the sample is calculated, using the previous relation. 
Finally, the bromide concentration of the sample is calculated 
using the relation 

[Br sample] = N sample (50.0 ppm / N std) (6) 

When a preliminary run revealed some interference from 
manganese (an isotope of which has an escape energy close to 
that of bromide-80) in the water, it was decided to count the 
longer lived isotope bromide-82 after waiting a day for the 
manganese-56 to decay. Because of this, it was necessary to 
make a correction to allow for the decay time prior to counting 
for all of the samples, including the standards. 

Six lysimeter nests were identified as representive of the 
range hydrologic and lithologic conditions found at the site. 
These were selected for intensive study. Their identification 
numbers, TR1LY1, TR1LY3, TR2LY1, TR2LY4, CLY5, and 
CLY6, are an artifact of the acid mine drainage abatement 
experiment, reflecting the existence of two treatment plots and 
one control plot. These nests contained a total of 23 functioning 
pressure-suction lysimeters buried at successive depths. A total 
of 257 neutron activation analyses have been successfully 
completed. This includes 17 ground water samples, 4 
background samples, 2 application samples and 234 soil water 
samples. This represents an average of over ten analyses per 
lysimeter. In addition, some sample activations were duplicated 
to ensure the validity of the results. 



INTERPRETATION OF RESULTS 

The use of concentration peaks to infer unsaturated 
hydraulic conductivities hinges on several assumptions. These 
can be divided into three categories pertaining to the behavior of 
the tracer, the integrity of the sampling program, and the 
behavior of the fluid flow in the reclaimed and revegetated strip 
mine. First, it is necessary to assume that bromide is a truly 
conservative tracer. With this in hand, we can assume that 
concentration peaks at a particular sampling point are 
proportional to average arrival times of tracer-bearing waters. 
Second, sampling protocol must be rigorous enough to prevent 
contamination of samples, which would lead to erroneous results. 
Also, the sampling schedule must be sufficiently dense to 
successfully define concentration peaks. In this case, the 
sampling schedule was limited not only by labor considerations, 
but also by the sampling devices themselves. Pressure-suction 
lysimeters collect water from the unsaturated zone at a 
decreasing rate with increasing unsaturation, as was the case in 
the early summer months of the study. However, the many 
well-defined peaks produced by this study suggest that the 
sampling schedule was sufficiently dense. Finally, for purposes 
of calculation of unsaturated hydraulic conductivity, it is 
necessary to assume that fluid flows in the shortest path from 
the land surface to the buried lysimeter. This assumption 
eliminates the evaluation of the nearly infinite variety of flow 



paths which infiltrating waters may have followed. Further, it is 
assumed that fluid flow occurs in a continuous and not discrete 
or pulsed manner. With a time scale on the order of days, this 
assumption may not be unreasonable. These two assumptions 
concerning the behavior of fluid flow will tend to skew the values 
obtained for unsaturated hydraulic conductivity downwards. 

With these assumptions in mind, the unsaturated 
hydraulic conductivites which exist around the pressure-suction 
lysimeter sampling points may be calculated by simply observing 
the time of the concentration peak at a particular lysimeter and 
dividing the depth at which the lysimeter is buried by this time. 
The unsaturated hydraulic conductivities presented in ascending 
order in table one are calculated based on the sampling date 
which yielded the highest concentration of bromide. 



Table 1. 


Computed unsaturated hydraulic conductivities. 


Lysimeter Lysimeter 


Lysimeter 


Unsaturated Hydraulic 


Nest 


Number 


Depth (ft) 


Conductivity (ft/day) 


CLY5 


74 


14.0 


0.0279 


CLY5 


73 


7.0 


0.0376 


TR1LY1 


24 


11.6 


0.0466 


CLY5 


75 


17.6 


0.0550 


TR1LY3 


34 


22.4 


0.0585 


CLY6 


78 


20.0 


0.0685 


TR2LY4 


56 


27.9 


0.0728 


TR2LY4 


54 


15.5 


0.0833 


CLY6 


76 


6.5 


0.1083 


TR2LY4 


55 


20.0 


0.1266 


TR1LY3 


32 


8.5 


0.1417 


TR2LY1 


83 


25.0 


0.1582 


TR1LY1 


23 


8.0 


0.1633 


CLY6 


79 


27.5 


0.1741 


TR2LY1 


43 


32.4 


0.1742 


TR1LY1 


27 


34.3 


0.2171 


TR2LY1 


84 


9.0 


0.2813 


TR1LY1 


25 


17.0 


0.3469 


TR2LY1 


82 


12.0 


0.3750 


TR2LY1 


42 


21.0 


0.4286 


TR2LY4 


57 


53.2 


0.4325 


TR2LY4 


53 


8.5 


0.5313 


TR1LY3 


35 


37.2 


1.1625 



It can be seen that these data range from 0.0279 to 0.5313 
ft/day unsaturated hydraulic conductivity with one case of 
1. 1625 ft/day. The bulk of the data fall in the lower end of the 
range. The high value almost certainly represents piping or 



FREQUENCY OF HYDRAULIC CONDUCTIVITIES 


N = 23 


11 <■ 


10' 






9- 






CO Q. 

c o 
•^ 1- 

O p.. 

£ 6 

cd 5- 

OT 

■° 4" 
O 

3- 






i 


2" 






■ 1 


1" 






ill ■ 


0" 


0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1.0 1.1 1.2 


K, FT/ DAY 


Figure 4. A frequency plot of the data from table 1. 



243 



channeling of infiltrating waters. Figure 4 is a frequency plot of 
the data from table 1. 

Several of the lysimeters exhibited concentration peak 
curves which had no rising or falling limbs. That is, instead of 
being smoothly increasing and decreasing, the curves had step- 
like jump discontinuities. This can be explained as a 
consequence of insufficient sampling frequency for the particular 
peak. However, the sampling frequency was sufficiently dense 
to capture the concentration peak itself in all cases. In two cases 
an initial concentration spike was separated by a period of six 
months from later bromide occurrence. In the case of lysimeter 
34, the two highest values, 1,161 ppb on April 14, 1987 and 
1,169 ppb on August 10, 1987, are not significantly different 
from each other. This evidence of unsaturated fluid flow 
behavior is somewhat problematic. One of us (Parizek) has 
suggested that this may be an artifact of a dual-porosity type 
mechanism. Here, fluid flow might be occurring on a variety of 
scales: interblock flow in voids between blocks of sandstones and 
in voids in shaly zones, flow in finer grained matrix materials, 
and intrablock flow in the pores of the rocks themselves. The 
capillarity of finer grained matrix materials and whole rock 
would tend to impede flow while the interblock voids would be 
much more conductive. If this model is correct, then it would be 
possible for a lysimeter to collect waters conducted through 
macropores very rapidly and through micropores much later. 

It is interesting to compare the values found in table 1 
with those yielded from pumping tests previously conducted 
(Henke 1985). These saturated hydraulic conductivities ranged 
from about 1 ft/day in coal tipple waste composed of about 50% 
clay and silt-sized particles to 100 to 110 ft/day in 
predominantly sandstone spoil. Other values included 8 ft/day in 
siltstone and shale, and 55 and 60 ft/day in sandstone spoil- 
centered pump tests. Excluding the high and low values from 
table 1, the mean and median unsaturated hydraulic 
conductivities found at the reclaimed and revegetated mine site 
are 0.1876 ft/day and 0.1582 ft/day, respectively. These are 
between 0.14 and 18.76 percent of the saturated hydraulic 
conductivities. 



Figure 6 shows the concentration values for all of the lysimeters 
installed in Control lysimeter nest 6. Non-ideal infiltration 
behavior is demonstrated here but is difficult to see because the 
magnitude of the concentrations varies over a wide range of 
values. 



BROMIDE VS. TIME 
CONTROL LYSIMETER NEST 6 



700001 
60000 
50000- 
40000" 
30000 
20000" 
10000- 





100 200 300 400 500 
Elapsed Time, DAYS 
Depth, FT -~ 6 . 5 



600 



20.0 < -*~S7.5 



Figure 6. A plot of Br concentrations versus time. 

Background concentration is below detection limits (20 
ppb). Application date is day zero. 



For this reason the concentration peaks for each lysimeter were 
normalized by dividing observed concentrations by the maximum 
observed concentration for that lysimeter. This simplified peak 
identification and interpretation. An example plot of these 
normalized values is shown in figure 7. 



Figure 5 is an example bromide concentration versus time 
plot for lysimeter 76 from Control lysimeter nest 6. The large 
primary peak is used for the calculation of unsaturated hydraulic 
conductivities. The exponential die-off curve suggests that 
storage capacities of the medium are not insignificant. 



700001 



60000' 



BROMIDE VS. TIME 
Lysimeter 76 




100 200 300 400 

Elapsed Time, DAYS 



500 



600 



Figure 5. A plot of Br concentrations versus time for a single 
lysimeter. Background concentration is below 
detection limits (20 ppbj. Application date is day zero. 



BROMIDE VS. TIME 
CONTROL LYSIMETER NEST 6 




100 200 300 400 

Elapsed Time, DAYS 



500 



600 



Deptti, FT i 



6.5 *-*-*20.0 *-*-T27.5 



Figure 7. An example plot of normalized Br concentrations 
versus time. Background concentration is below 
detection limits (20 ppb). Application date is day zero. 



Here, we can clearly see the non-ideal behavior, with the first 
peak in the shallow lysimeter, the second peak in the deep 
lysimeter, and the final peak in the intermediate depth 
lysimeter. 



244 



Figure 8 shows the apparent randomness with respect to 
depth of the hydraulic conductivities. This is consistent with the 
type of lithologic configurations one would expect to find in 
disturbed mine spoil. The vertical alignment of unsaturated 
hydraulic conductivities may be an artifact of insufficient 
sampling density, or it may be because of similarities in 
lithologies found at these various sampling points. 



LYSIMETER DEPTH 

VS. 

CONCENTRATION PEAK ARRIVAL TIME 



601 
50 
£40 

^30 

20 

10 





Q. 
CD 
Q 





C 


Symbols Correspond 
To Treatment Plots 


2 

C 
2 

0S22 


2 C 

1 

c 
c 

1 


C 

2 1 



100 200 300 400 500 600 
Peak Arrival Time, DAY 



Figure 8. Scatter plot of peak arrival times vs depth. Points 
labelled 1, 2, and C are from treatments 1, 2, and the 
control plot, respectively. Rather than early peaks 
occurring at shallow depths, peak arrivals appear to 
be independent of depth. This is indicative of non- 
uniform, non-ideal infiltration and percolation behavior 
in the unsaturated zone. 



With the peak arrival data in hand, and keeping in mind the 
assumptions previously mentioned, it is possible to calculate 
unsaturated hydraulic conductivity. Dividing depth by time 
yields the numbers depicted in figure 9, a scatter plot of depth 
versus unsaturated hydraulic conductivity. An examination of 
figure 9 suggests an irregular hydrogeologic regime, where areas 
of high and low hydraulic conductivity do not occur in distinct 
horizons, but rather are distributed more randomly with respect 
to depth. This insight helps us to build a more complete 
conceptual framework and understanding in regards to fluid flow 
in disturbed land. 

With figures 8 and 9 in mind, table 2 is a listing of the 
lithologies of the lysimeter sampling points in the same order of 
increasing unsaturated hydraulic conductivity used in table one. 

Table 2 describes a highly irregular backfill spoil pile consisting 
primarily of two sedimentary rock types, sandstones and shales. 
Perhaps more significant than the lithological descriptions in 
table 2 is the lack of descriptions. In many cases the log reports 
no return or poor return and bit drops. Furthermore, the process 
of backfilling sometimes required quantities of sediment many 
times the borehole volume to complete. All of this evidence is 
testimony to the frequent presence of large void spaces in the 
spoil pile. 



ONGOING AND FUTURE STUDIES 

In terms of the unsaturated fluid flow characterization, the 
determination of soil moisture content and bulk density are 
critical. Volumetric water content is the ratio of volume of water 



LYSIMETER DEPTH 

VS. 

UNSATURATED HYDRAULIC CONDUCTIVITY 



601 



50- 



£40 



CL 
Q) 

Q 



30 



20 



10 



Symbols Correspond 
To Treatment Plots 



C 2 

l c 

2 



2 

c 



0.0 0.2 0.4 0.6 0.8 1.0 1.2 
Hydraulic Conductivity, FT/DAY 



Figure 9. Scatter plot of unsaturated hydraulic conductivities 
versus depth. Points labelled 1, 2, and C are from 
treatments 1, 2, and the control plot, respectively. 
Randomly oriented values provide insight into the 
nature of fluid flow in disturbed mine spoil. 



Table 2. Lithologies of the lysimeters. 



Lysimeter 
Number 



Lithology 



74 Grayish-br sand and sandstone and shale/siltst frags 
73 Grayish-br sand and sandstone and shale/siltst frags 
24 Orangish-br sand and sandstone frags 

75 Grayish-br sand and sandstone and shale/siltst frags 
34 Poor return, bit drop® 10 ft, boulder@27 ft 

78 Poor return, gray to yellowish-br sand 
and sandstone frags, some coal frags 

56 No return, bit dropped 0.5 ft@5,15 and 20 ft 

54 No return, bit dropped 0.5 ft@5,15 and 20 ft 

76 Gray to yellowish-br sand and sandstone frags, 
some dark-gray shale/siltstone and coal frags 

55 No return, bit dropped 0.5 ft@5,15 and 20 ft 
32 Gray to brownish-gray sandstone fragments 

83 Poor return, gray sand and sandstone frags and 
dark-gray shale frags 

23 Gray sandstone frags and yellowish-br sand 

79 No return 

43 Poor return, gray sand and sandstone frags and 

dark-gray shale frags 
27 Poor return, br sand and sandstone frags 

84 Poor return, gray sand and sandstone frags and 
dark-gray shale frags 

25 Orangish-br sand and sandstone frags 

82 Poor return, gray sand and sandstone frags and 

dark-gray shale frags 
42 Poor return, gray sand and sandstone frags and 

dark-gray shale frags 
57 No return, bit dropped 0.5 ft@5,15 and 20 ft 
53 No return, bit dropped 0.5 ft@5,15 and 20 ft 
35 Poor return, bit drop® 10 ft, boulder@27 ft 



245 



to local volume, while bulk density is the ratio of dry mass of 
solid to total mass (Fritton 1986). Recalling that unsaturated 
hydraulic conductivity is a function of volumetric moisture 
content, this parameter is of interest. Knowing volumetric 
moisture content and bulk density, total porosity of the 
conducting medium can be calculated. An understanding of total 
porosity and bulk density will help to gain insight into the nature 
of the matric materials and their arrangement, which have a 
Luge influence on matric potential. Although the tracer study 
has been underway for some time, it is still desirable to obtain 
some representative values for soil moisture content under 
different meteorological conditions. This enhancement of the 
breadth and scope of the unsaturated fluid flow characterization 
of mine spoil is the motivation for the depth-density and depth- 
moisture profiling that is in progress. The specific relation 
between bulk density, total porosity, and volumetric moisture 
content is 



where 



E = 1 - (6 vPw /p s - p t /p s ) (3) 

E = Total Porosity 

p t = Bulk Density, M t /V t 

B v = Volumetric Water Content, V w /Vt 



The values for By and pi can be empirically determined from in 
situ investigations using depth-moisture and depth-density 
neutron thermalization probes. p w is taken to be 1.0 g/cm 3 and 
Ps is taken to be 2.65 g/cm 3 -. This derivation is based on the 
relations of Fritton (1986). 

With the goal of depth-moisture and depth-density profiling 
in mind, six 20 ft-long very thin walled (0.035 in.^ seamless 
aluminum access tubes were installed at the site in early 
September 1987. The installation locations were chosen to 
investigate the several characteristic heterogeneities and 
anisotropies found at the site. Further, the access tubes were 
installed proximal to the lysimeter nests whose flow behavior 
has been investigated in detail through large numbers of neutron 
activation analyses. With this strategy, a more definitive and 
detailed understanding of the fluid flow behavior at the site may 
be realized. 

In addition, we recognize that sampling schedules may 
produce artificially skewed concentration peak dates and thus 
alter calculated unsaturated hydraulic conductivity. In the light 
of this, a remedial solution may be found in geostatistical 
approaches. That is, the determination of a semivariogram 
containing the concentration values at the sampling point 
through time may yield somewhat different peak dates. 



micaceous, sandstones and finer grained disaggregated shales. 
The presence of many void spaces in these disturbed sediments 
is suggested both by the lithologic logs of the drill holes and by 
the difficult experience backfilling some of these same holes, 
which often required many times the borehole volume to 
complete. Despite these void spaces, only one suspected case of 
chanelling was found. Consistent with other undisturbed 
environments, in the disturbed mine spoil unsaturated hydraulic 
conductivity is one to three orders of magnitude less than 
saturated hydraulic conductivity. The values calculated have a 
narrow range and appear to be largely independent of localized 
lithologic heterogeneities. This, however, may be an artifact of a 
sketchy drill log record. Accurate and complete drill logs from 
disturbed mine spoil materials continue to prove difficult to 
obtain, as cuttings are often lost to the subsurface voids. With 
these unsaturated hydraulic conductivity values in hand, it is 
possible to calculate volumetric and mass flux of percolating 
waters through a unit cross-sectional area under a unit hydraulic 
gradient. Given an estimate of alkaline load in percolating soil 
water, either empirical or theoretical, it is possible to calculate 
acid neutralization potentials for a given mass or volume of 
water. Comparing these figures with acid production estimates 
should yield valuable insight into the effectiveness of a given 
surficial application of alkaline material. Clearly, the dramatic 
loss of hydraulic conductivity under unsaturated conditions 
presents a significant retardation factor in the speed and 
effectiveness of in situ abatement efforts. 

Further refinement of the data may be yielded by 
geostatistical investigations, as the semivariogram of the 
concentration values at a sampling point through time may 
suggest alternative concentration peak dates. A second 
investigative technique, the use of neutron depth-density and 
depth-moisture probes, will be of use in the determination of soil 
bulk density and soil moisture content values. These data will 
help to quantify the range of unsaturation under which these 
spoil waters moved. 

The abatement of acid mine drainage from strip mines 
through the use of surficial applications of lime plant flue dust 
and limestone quarry waste depends heavily on alkaline-charged 
waters percolating through acid-producing spoil material. Thus, 
an increased understanding of the behavior and transport rates 
of water through the unsaturated zone is of great value. With 
the unsaturated hydraulic conductivity values yielded from this 
study, investigators will be better prepared to evaluate past 
abatement strategies and plan future abatement strategies. 



LITERATURE CITED 



CONCLUSIONS 

A tracer study using non-radioactive neutron activatable 
bromide-79 was employed as an investigative technique to 
determine in situ unsaturated hydraulic conductivities in a 
reclaimed and revegetated strip mine. Subsequent to the 
application, water samples were collected through time from 55 
of the pressure-suction lysimeters installed at various depths at 
the field area. 23 were selected for study and neutron activation 
analysis was used to determine Br concentration. Interpretation 
of the concentration peaks as average arrival times of infiltrating 
rainwater yielded unsaturated hydraulic conductivities ranging 
from 0.0279 to 0.5313 ft/day with one value of 1.1625 ft/day. 
This high value is interpreted as an example of infiltrating water 
piping through the highly heterogeneous disturbed spoil pile. 
Excluding the high and low values, the mean and the median 
unsaturated hydraulic conductivities were found to be 0.1876 
ft/day and 0.1582 ft/day, respectively. These values are in the 
range of 0.14 to 18.76 percent of saturated hydraulic 
conductivity of the various geologic materials at the site. 

The spoil at the site is composed of blocky, often 



Brasino, J. A. and J. A. Hoopes. 1985. Comparison of 
Unsaturated Zone Movement Between Tracer Potassium 
Bromide and Pesticide Aldicarb. In: Proceedings of the 
NWWA Conference on Characterization and Monitoring of the 
Vadose (Unsaturated) Zone, November 19-21, 1985. The 
National Water Well Association, Dublin, OH. pp. 347-254. 

Freeze, R.A. and J.A. Cherry. 1979. Groundwater. Prentice- 
Hall, Englewood Cliffs, NJ. 604 pp. 

Fritton, D.D. 1986. Unpublished Class Handout, Agronomy 
401, The Pennsylvania State University, University Park, PA. 

Henke, J.R., 1985. Hydrogeologic Characterization Of A Surface 
Mining-Impacted Watershed with Implications for Acid Mine 
Drainage Abatement, Clarion County, Pennsylvania. The 
Pennsylvania State University, University Park, PA. 
Unpublished Master's Thesis, 171 pp. 

Jester, W.A., D.C. Raupach and F.G. Haasar. 1977. A 
Comparison of the Bromide Ion Neutron Activatable Tracer 
With Tritiated Water as Groundwater Tracers. In proceedings 
of the 3 rd Int. Conf. Nuc. Meth. Env. & Energy Res. 
CONF-771-072. pp. 253-264. 



246 



Miller, F.W., Jr. 1982. College Physics. Harcourt, Brace, 
Jovanovich, Inc. New York. 876 pp. 

Raupach, D.C., W.A. Jester, E.T. Shuster and C. Brethauer. 
1983. Using A Neutron Activatable Tracer To Indentify An 
Acid Mine Drainage Source. Paper presented at the meeting of 
the American Nuclear Society, Detroit, ML, Nuclear 
Engineering Department, The Pennsylvania State University, 
University Park, PA. and Bureau of Water Quality 
Management, Pennsylvania Department of Environmental 
Resources. 5 pp. 

Schmotzer, J.K., W.A. Jester and R.R. Parizek. 1973. 
Groundwater Tracing With Post Sampling Activation Analysis. 
Journal of Hydrology 20: 217-236. 



247 



FORECASTING THE EFFECT OF MINE SITE REHABILITATION 
WORKS ON LOCAL GROUND WATER QUALITY 1 



David K. Gibson and Garry Pantelis 2 



Abstract.—A large rehabilitation project has been carried out on an 
abandoned uranium mine at Rum Jungle in the Northern Territory of 
Australia where oxidation of pyritic mine wastes has led to substantial water 
pollution. This pyritic oxidation has been halted by the exclusion of air and 
water from two rock heaps by contouring and installing a clay cap. Ground 
water levels and quality have been monitored for acidity and dissolved metals 
over the four years since the project began. Although the contaminant input 
to the local river has dropped markedly, there has been no significant change 
in ground water quality. The purpose of this work is to estimate the time 
scale over which an improvement might be expected to occur. A model in 
which the main store of pollutant is assumed to be the contaminated water 
below the heap leads to the very low estimate of less than four years. A more 
complex model which accounts for contaminants stored in pore water held 
within the dump produces an estimate, so far consistent with observation, of 
20 years. The possibility of heavy metals being stored in the form of a 
buffered precipitate cannot be excluded. Thus, the importance of continuing 
the monitoring of the site is emphasised. 



INTRODUCTION 

Rum Jungle is on the East Finniss River, about 70 km 
south of Darwin in the Northern Territory of Australia. From 
1954 to 1971 it was the site of a mining operations where 
uranium and copper were extracted by opencut mining. The 
East Finniss River was significantly contaminated by heavy 
metals during mining, and this continued after the mine was 
abandoned. The mean concentrations of the metals in the river 
were a few mg/1, but owing to the strong seasonal variation in 
the rainfall, the actual concentrations varied widely, being 
rather high at the periods of small flow at the beginning and 
end of the wet season. The average rainfall is about 1.5 m/y 
falling mainly in a wet season over the period October to 
April. 



1 Paper presented at the 1988 Mine Drainage and Surface 
Mine Reclamation Conference, sponsored by the American 
Society for Surface Mining and Reclamation and the U.S. 
Department of the Interior (Bureau of Mines and Office of 
Surface Mining Reclamation and Enforcement), Pittsburgh, 
Pa, April 17-22, 1988. 

2 David K. Gibson is a Senior Research Scientist and Garry 
Pantelis is a Research Scientist, Physics of Environment 
Section, Environmental Science Division, Australian Nucle- 
ar Science and Technology Organisation, PMB 1, Menai, 
NSW 2234, Australia. 



During the 1970s the environmental problems of the site 
were investigated to elucidate the nature of pollution caused by 
mining. It was demonstrated that most of the heavy metal 
pollution arose from three large overburden dumps (one of 7 
million tonnes and two of 2 million tonnes) and a heap leach 
pile, in which pyrite was oxidising, catalysed by bacteriologi- 
cal activity (Harries and Ritchie 1987). Heavy metals in the 
pyritic rock were solubilised in this process and were eventual- 
ly transported by water movement into the river system. 

The rehabilitation works were carried out from 1983 to 
1985. The tailings were disposed of in one of the three 
opencut pits. The water in the other two pits was treated, the 
pH being raised from about 3.5 to 7.5, with consequent 
precipitation of the dissolved heavy metals. The tops of the 
waste rock heaps were contoured and capped with layers of 
clay, loam, and sandy gravel with total thickness of about 600 
mm (Northern Territory Department of Mines and Energy 
1986(a)). The dumps were vegetated with a mixture of grasses 
and legumes, and erosion was controlled with a system of 
bunds and rock drains. The sides of the heaps were sloped at a 
maximum of 1 in 3 and faced with rock. 

The clay cover was designed to reduce the infiltration of 
water through the heaps. Measurements with collection 
lysimeters have shown that the infiltration of rain water has 
been reduced to less than 10% of its former value. Measure- 
ments of the oxygen levels and heat generation in the dumps 



248 



have shown the pyritic oxidation has also been greatly reduced 
(Harries and Ritchie 1987). Two overburden heaps were 
selected for intensive study. The ground water in their vicinity 
was monitored by about 30 boreholes for the purpose of 
estimating the rate at which the quality of the ground and 
surface water could be expected to improve over the coming 
years following the cessation of pollutant generation. 

MODEL OF POLLUTANT RELEASE FROM BELOW 
WASTE ROCK HEAP 

Estimated pollutant release was based on a simple 
two-dimensional model applied to the White's heap, the 
largest waste rock heap. This heap extends for about 600 m 
along a gently sloping ridge. It is about 20 m high and 600 m 
wide. The ridge continues for some 200 m upslope of the 
dump to a water divide at the top of the hill. The underlying 
ridge has a permeable layer of soil about 2 m thick sitting on 
fairly impermeable rock. This is shown in schematic form in 
figure 1, in which the x-axis coincides with the fall of the 
ridge. 

The water in the aquifer below the dump is a reservoir of 
dissolved heavy metals, generated before the clay cap was 
installed. This contaminated water will be flushed out as 
rainfall recharges the aquifer upslope of the heap. An estimate 
of the time taken for this flushing process should provide a 
lower bound for the time before any improvement in ground 
water quality can be expected as other processes, such as the 
flushing of contaminated water stored in the unsaturated zone 
in the dump itself, will add to this time. 

Assuming steady-state conditions, we can calculate a 
water table height by using the Dupuit-Forscheimer equation 
(Bear 1972). If the phreatic surface is represented by (|)(x), the 
base of the aquifer by g(x), the saturated hydraulic conductivi- 
ty by K t , and the total horizontal discharge by q, we have 



dx 



d_ 
'dx 



dh 



= 0, 



0<x<L 



(1) 



where Q is the recharge from the upslope aquifer. From (1) 
and (2) we simply have for steady conditions 

q(x) = Q 0<x<L (3) 

If we take the convection-dispersive equation (CDE) for 
one dimensional saturated horizontal flow, assume uniform 
vertical concentration and ignore dispersion, we have 

dt hdx 

where h = <j)-g, B s is the saturated moisture content of the 
aquifer and c is the contaminant concentration. Under pure 
advection the fresh water/contaminated water interface is 
given by the characteristic curve of (4), which is 



dx 

dt 



_Q_ 



'BJi 



(5) 



emanating from x = at time t = 0. 



Since h is a function of x we may invert (5) and solve for t as a 
function of x, i.e. 



dt(x) QMx) 



dx 



Q 



(6) 



The travel time, t , for contaminants starting at x = (t=0) and 
reaching x=L is obtained from 



6, C L 
'o = 77 h{x)dx 
V. Jo 



Noting that here <h(x) < 2m we have 

2Q S L 



0<t o ^- 



Q 



(7) 



(8) 



At x = we have the boundary condition 



(2) 




Figure 1.— Schematic diagram of mine waste overburden 
heap lying on sloping ridge. The vertical scale is 
greatly exaggerated. 



An estimated value of Q is given by the product of the 
length of the recharge region (200 m), the annual rainfall (1.5 
m/y) and the infiltration coefficient (0.5). Thus Q = 150 m 2 /y 
and an estimate of 0, = 0.4 were used to yield the estimate of 
< t o < 3.2 years for the flushing-out period. 

A more elaborate model has also been used for this 
problem, in which the flushing of the unsaturated moisture 
from the heap was included (Pantelis 1987). A vertical 
averaging procedure was used to deal with the contaminant 
profile. In order to avoid making the assumption of uniform 
vertical concentration in the CDE the product of moisture 
content and contaminant concentration, 8c, was replaced by its 
vertical average. This somewhat artificial manipulation intro- 
duced the vertical integral of 6c as the dependent variable. 
This approach implies a more vertical uniformity of contami- 
nant mass rather than of concentration, which may be more 
realistic since, in the absence of recharge at the top of the 
overburden heaps, low moisture contents and low mobility of 
contaminants could result in higher concentrations existing 
there. A high concentration of contaminants in the pore water 
in the upper levels of the dump is supported by measurements 
made on samples taken from the heap (Goodman et al. 1981). 
However, this assumption, together with the low percentage of 
rainfall infiltrating the clay capping, could lead to an overesti- 
mate of the flushing time. A period of 20 years was evaluated 
by this model, which, for the above reasons, must be taken as 
an upper limit. It is stressed that the flushing time could be 
longer if heavy metals continue to be mobilised, for instance 
by the remobilisation of metals deposited as minerals during 
the period of pyritic oxidation. 



249 



FIELD MEASUREMENTS 

Around the time of the capping of White's heap about 20 
boreholes were drilled into the aquifer at various points about 
the perimeter of the heap; one hole was drilled through the 
heap into the aquifer below. These have been monitored for 
water level, pH, electrical conductivity, and concentrations of 
copper, zinc, manganese and sulphate, approximately at 
monthly intervals, from the end of 1983 till the present. 
Results for a typical borehole near the perimeter of the heap 
and for the borehole through the heap are shown in figures 2 
and 3. Estimates of the total annual contaminant loads in the 
Finniss River have also been made (Northern Territory Depart- 
ment of Mines and Energy, 1986(b)) and are shown in table 1. 



As can be seen from figure 2, the ground water quality 
has not changed significantly over the four years since 
rehabilitation. It would take considerable stretching of the 
parameter values used in the simple model to achieve consis- 
tency with this time. It is therefore evident that there is a store 
of contaminants additional to that in the ground water. The 
20- year period estimated using our second model, which 
accounts for the storage of contaminants in pore water held in 
the dump, is consistent with the data, as is the prediction that 
no change is expected for the period, after which change is 
rapid. However, it cannot as yet be demonstrated whether this 
store is the controlling factor or whether the contaminants are 
stored in the form of a buffered precipitate. 



69.0 



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i*0 

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80 
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60 
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85 86 

Cu Cmg/I 3 



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95 



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18 






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8k 85 86 87 

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Electrical conduct I v\ ty CmS/cm3 
3k I 1 1 1 1 



_L 



85 86 87 

Borehol e posl t\ on 




8k 85 86 87 

Figure 2.~Data from borehole No. 22084 from 1984 to 1987. 



250 



It is interesting to note the seasonal variation of water 
level and contaminant concentration observed beneath the 
heap and shown in figure 3. There is obviously considerable 
water movement below the dump. It also appears that the 
contaminant concentrations fall during the wet season, due to 
dilution by fresh water, and then rise during the dry. This 
could be due to the redissolving of precipitates or to input from 
the concentrated pore water. 

Table 1 shows that the total input of heavy metals to the 
river system has fallen markedly. Again this is consistent with 
the prediction of the more complex model that the total 
discharge of pollutants from waste rock dumps will decrease 
after capping, because of reduced water discharge. 



CONCLUSIONS 
The field measurements show that it takes at least four 
years before ground water quality in the vicinity of a large 
overburden rock dump will respond to measures that stop 
leaching of heavy metals from it. We have shown that this 
period cannot be accounted for on the assumption that the 
main store of contaminants is in the ground water below the 
dump. It is, however, consistent with a more complicated 
model in which contaminants are assumed to be held in 
concentrated solution in the pore water held in the rock heap. 
As the system under investigation is a major rehabilitation 
work and has been well monitored from the outset, it is most 
important that monitoring continue until the trend in ground 
water quality becomes clear. 



Uot er level ImJ 



72 

71 
70 

69 






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30 
20 
10 



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h 




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Electrical conductivity CmS/cm] 



35 - 
30 : 
25 
20 






8^ 85 86 87 

Boreho I e position 




8^ 



85 



86 87 



Figure 3.-Data from borehole No. 22082 from 1984 to 1987. 



251 



Table 1 .-Summary of monitoring results for the East Branch of the Finniss River. 



Season 


1982/83 


1983/84 


1984/85 


1985/86 


Rainfall (mm) 


1121 


1704 


1112 


910 


Total flow (m'xlO 6 ) 


9.5 


48 


11.7 


11.4 


Metal load (t) 
Copper 


23 


28 


9 


4 


Manganese 


5 


9 


4 


3 i 


Zinc 


5 


9 


4 


3 



LITERATURE CITED 

Bear, J. 1972. Dynamics of fluids in porous media. American 
Elsevier Publishing Company, N.Y. 

Goodman, A.E., Khalid, A.M. and Ralph, B.J 1981. Microbial 
Ecology of Rum Jungle Part 1. Environmental Study of 
sulphidic overburden dumps, experimental heap-leach 
piles and tailings dam area. Australian Atomic Energy 
Commission/E53 1 . 

Harries, J.R. and Ritchie, A.I.M. 1987. 

Environmental Geochemistry and Health. The effect of 
rehabilitation on the rate of oxidation of pyrite in a mine 
waste rock dump 9, 2, 27. 



Northern Territory Department of Mines and Energy, 1986(a). 
The Rum Jungle Rehabilitation Project: Final Project 
Report. Northern Territory Department of Mines and 
Energy, Darwin, Australia. 

Northern Territory Department of Mines and Energy, Water 
Resources Division 1986(b). Summary of water re- 
sources monitoring at Rum Jungle December 
1982-February 1986. 

Pantelis, G. 1987. Modelling water and contaminant transport 
in the Rum Jungle mine overburden heaps, Australian 
Atomic Energy Commission, AAEC/E653. 



252 



PHOTO-LINEAR CHARACTERIZATION, LITHOLOGIC VARIABILITY, AND THE EFFECTS OF 
MINING ACTIVITY. BY FRACTURE STUDIES AND IN SITU, AIR-INJECTION, 

PERMEABILITY TESTING * 



Christopher A. Shuman and Richard R. Parizek 2 



ABSTRACT 

Highwall fracture studies and a novel permeability 
testing technique were used in a study of the 
subsurface character of a photo-linear "fracture 
trace". This study focused on a topographic-tonal, 
photo-linear feature that intersected the highwall of 
an active surface coal mine working the Freeport- 
Mahoning coals near Clearfield. Aerial photographs 
at 1:12,000 scale were used for the photo-linear 
analysis, and critical features were examined in the 
field. Fracture data on joints and faults were 
gathered in traverses along highwall exposures of a 
relatively uniform siltstone unit overlying the Upper 
Freeport Coal, and hydrologic tests through borehole 
packers were conducted in overburden shotholes and 
selected monitoring boreholes drilled into the 
siltstone. The orientation of photo-linear and joint 
distribution maximas compare well, both falling 
between 320 - 340 degrees. Joint frequency analysis 
indicated that fracturing increased slightly near the 
photo-linear feature's subsurface projection. 
However, the correlation of this feature with a swale 
in the Mahoning Coal (possibly an old distributary 
channel) indicated that fracture frequency increases 
are probably related to stratigraphic phenomena 
rather than "fracture traces". The air-based packer 
test technique used for the permeability testing was 
designed to gather data rapidly from available 
boreholes, especially those in unsaturated units 
adjacent to the mine pit. A total of 55 permeability 
values were obtained from the siltstone. These 
values defined the permeability variation within the 
siltstone (probably 1.0 to 7.0 x 10" 13 ft*), 
evaluated the effect of mining activity on bedrock 
permeability ( 2x + increases extending 20-30 ft from 
the highwall), and enabled the air-permeability 
values to be correlated with the hydraulic 
conductivity values (3.9 to 5.1 x 10 - 7 ft/sec) 
derived from water-based tests. In addition, 
air-injection test factors (such as Klinkenberg 
slippage, partial saturation, and turbulent flow) and 
some additional uses of the technique were given a 
preliminary evaluation. 



»Paper given at the 1988 Mine Drainage 
and Surface Mine Reclamation Conference 
sponsored by the American Society for 
Surface Mining and Reclamation and the 
U.S. Department of the Interior (Bureau of 
Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 
1988, Pittsburgh, PA. 



2 Christopher A. Shuman is a Research 
Assistant and Richard R. Parizek is a 
Professor of Geology, The Pennsylvania 
State University, University Park, PA. 



253 



INTRODUCTION 

One topic which has recently received 
more than passing attention from hydro- 
geologists is the movement of ground water 
through subsurface fractures. This 
interest is due to the ubiquitous nature 
of geologic fractures, their generally 
permeable character, and the fact that 
they can interfere with the application of 
Darcy's Law. By studying the nature of 
fracture flow, as well as the distribution 
and character of fracture features, 
ground-water scientists can better adapt 
porous media theory, based on Darcy's Law, 
to fit particular geologic situations. 
This will improve understanding and 
prediction of ground-water flow in many 
geologic systems. However, a great deal 
of information on the subsurface nature of 
fracture features and their effects on 
porous media must be gathered before 
consistent evaluation of these systems 
will be generally possible. 

The primary objective of this study 
was a quantified evaluation of the 
geologic and hydrologic character of the 
subsurface expression of a photo-linear 
"fracture trace" feature. Because these 
features have been correlated with zones 
of weakness and increased ground-water 
flow ( Lattman and Matzke 1961, Parizek 
1976), it was expected that this research 
would yield valuable field data on the 
nature of these features that could be 
applied in other areas. This was to be 
achieved by fracture studies and in situ 
permeability tests that could be focused 
where a mapped feature intersected the 
highwall of a surface coal "strip" mine. 
Secondary objectives to be studied 
included the variability of permeability 
within a fractured lithology, the effects 
of mining on permeability, and the factors 
affecting the field use of air-injection 
permeability tests. 

This study was conducted in Clearfield 
County in the Appalachian Plateau region 
of Pennsylvania. The Browncrest III 
surface coal mine, operated by the Central 
Pennsylvania Coal Company of Clearfield, 
was selected as the study site because of 
the availability of recent aerial 
photograph coverage of the mine, the 
presence of a photo-linear feature that 
intersected the mine pit, the relatively 
uniform and undisturbed nature of the 
exposed lithologies, as well as the 
cooperative attitude of the mine operator. 

RESEARCH TECHNIQUE 

As illustrated by figure 1 , the 
research technique employed in this study 
had three phases: 1) identification of a 
photo-linear fracture trace feature where 
it crossed a mine highwall, 2) measurement 
of joints and faults in traverses of a 
suitable highwall lithology near the 
fracture trace and in other areas of the 
pit, and 3) development and application of 
the permeability testing technique in 
available overburden shotholes and 
monitoring boreholes drilled in the 
vicinity of the feature. By examination 




Figure 1 -- Location map of study area 
features . 

of the exposed lithologies (from the Lower 
Freeport Coal to a sandstone above the 
Mahoning Coal), it was decided to 
concentrate the fracture studies and 
hydrologic testing on a siltstone unit 
that overlies the Upper Freeport coal (see 
figure 2). This unit was selected because 
of the number of drillholes into it, as 
well as its consistent character and 
exposure across the mine. 

Photo-linear Fracture Trace Analysis 
The fracture trace analysis was done 
in order to characterize the overall 
distribution of photo-linear features and 
to provide a basis for comparision with 
other available data on joints and 
fracture traces for this region. The 
technique used to map the photo-linear 
fracture trace features at the mine site 





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Upper Freeport (E) Coal 







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sandstone 

shale 

siltstone 

claystone 

mudstone 

limestone 

coal stringer 

coal 

undercloy 




























































m 




w 







Freeport Rider (DR) Coal 



Lower Freeport (D) Coal 



VERTICAL SCALE 

0~ ~ , 30 

Feet 



Figure 2 -- Composite stratigraphic 
section for the study area. 



254 






was derived from Lattman (1958) and Meiser 
and Earl (1982). Aerial photographs at 
1:12,000 scale were studied individually 
and in stereo with the aid of a mirror 
stereoscope. The ends of observed 
features were marked on acetate overlays 
and the fracture trace feature shown in 
figure 1 was then field-checked for 
accuracy. Previous mining activity to the 
south and east of the mine pit made the 
analysis difficult. However, the three 
images which provided the photographic 
base for this analysis were of excellent 
quality, revealed land surface details 
with great clarity, and were only a few 
months old when this study began. 
Highwall Fracture Studies 

In order to determine if the photo- 
linear fracture trace feature intersecting 
the mine pit was related to a subsurface 
zone of fracture concentration, highwall 
fracture studies were conducted. The 
objective of this phase of the research 
was a qualitative and quantitative 
evaluation of joints and other geologic 
fractures that could be observed in 
highwall traverses. Joint measurements 
were made in 3 traverses, in the vicinity 
of the mapped feature as well as in areas 
thought not to be affected by linear 
features. Joint and fault planes were 
distinguished from blasting related 
fractures on the basis of their extent, 
regularity, degree of weathering, and 
relationship to remanents of shotholes 
visible along the highwall. The mapped 
joint and fault features were located 
along the 3 traverses with a fiberglass 
measuring tape. A Brunton compass, scale, 
and set of descriptive criteria were used 
to characterize each joint or fault 
identified along a traverse. 

Air-Injection Permeability Testing 

In order to conduct the density of 
permeability tests necessary to quantify 
the subsurface nature of the mapped 
photo-linear fracture trace, a novel 
air-injection permeability measurement 
technique had to be developed. This 
technique, illustrated in figure 3, was 
designed to quickly and accurately gather 
bedrock permeability values from test 
intervals at the bottom of overburden 
shotholes drilled in the siltstone. By 
testing in the patterns of these shotholes 
(closely spaced boreholes with similar 
characteristics), information on the 
spatial variation and factors affecting in 
situ permeability determinations was 
gathered . 

Due to the nature of the surface mine 
environment, the technique had to be 
mobile, durable, and designed to operate 
in unsaturated conditions. The equipment 
item which made this possible was the 
borehole packer. By using a single packer 
to isolate the bottom of the borehole, 
permeability values were obtained from the 
constant-head (steady-state) air-pressure 
and corresponding air-flow conditions that 
could be developed in each test interval. 
These air tests were patterned on 
constant-head water tests used to gather 
hydraulic conductivity data under 
saturated conditions (Houlsby 1976). 



HOIST TRUCK 
/ 



AIR-INJECTION 
EQUIPMENT TRAILER 




Figure 3 — Schematic of air-injection 
permeability test. 

Because of the severe limits on testing 
time imposed by the active mining, as well 
as equipment and manpower restrictions, it 
was not possible to gather subsurface 
fracture data from tested intervals with 
an impression packer. 

The standard testing format, following 
insertion and inflation of the single 
packer, was a multiple step (abcba) test 
conducted in a sequence of pressures of 
approximately 10, 20, 30, 20, and 10 
lb/in 2 (psi). For each injection 
pressure step, a flow rate in ft 3 /min 
(cfm) was recorded. Following completion 
of the last pressure step, the packer was 
deflated, withdrawn, and the testing 
apparatus was moved to the next shothole. 
All data were analyzed by the equation: 

k=[uQ/[2nh(Pw-Pa)]] ln(r e /r„) 
where k is material permeability, u is 
dynamic viscosity, Q is fluid flow rate, 
n is pi, h is tested length of bore- 
hole, Pw is average borehole pressure, Pa 
is atmospheric pressure, r» is the 
effective borehole radius, and r w is the 
actual borehole radius (Blankenship and 
Stickney 1983). The multiple-pressure 
step method was designed to yield data on 
the behavior of test factors such as 
Klinkenberg slippage, turbulent flow, 
partial saturation, and other complicating 
factors (Klinkenberg 1941, Pearson and 
Money 1977) . 

The tests were usually run in the 
overburden shotholes (20 to 40 holes in a 
pattern) in the time between drilling and 
detonation. This restriction prevented 
all retesting of the boreholes as well as 
borehole development. However, the method 
proved capable of providing numerous 
permeability values from the siltstone 



255 



unit (average borehole spacing of 
approximately 15 ft). Other tests were 
conducted in monitoring boreholes outside 
the immediate mining area. Unfortunately, 
these holes were not as close to the mine 
highwall and were partially saturated with 
water which made them unsuitable for 
air-testing. They were suitable for the 
more traditional hydrologic tests which 
were conducted as a check of the 
air-injection technique's accuracy. 
Other Hydrologic Tests 
In order to obtain hydraulic conduc- 
tivity values from the siltstone unit 
which could then be compared to the 
air-based permeability data, both 
falling-head (slug) and constant-head 
'steady-state) water injection tests, 
using standard methods and analysis 
routines, were conducted in the available 
monitoring boreholes (Fetter 1980, Cooper 
et al 1967, Price et al 1982). These 
boreholes served as the site for tests of 
other applications of the air-injection 
system, such as falling-head air-injection 
tests and air-injection interference tests 
between adjacent boreholes. 

RESULTS AND DISCUSSION 

Photo-Linear Fracture Trace Analysis 
The results of this phase of the 
research are presented in figure 4 which 
shows the orientation diagram for the 24 
features observed in the vicinity of the 
mine site. These features showed a broad 
orientation maxima between 300 and 340 
degrees of azimuth with a peak between 320 
and 330 degrees. The features ranged in 
length from 2,000 to 5,500 ft and were 
defined by alignments of topographic, 
tonal, and or vegetative features. The 
photo-linear feature of interest to this 
study was identified by a combination 
alignment of topographic and tonal 
elements trending to the northeast of the 
mine. The orientation results compare 
well to other published results (Lattman 
and Nickelsen 1958) and the distinct 
orientation of several local tributary 
streams . 

Highwall Fracture Studies 
Prior to the fracture mapping of 
traverses along the siltstone unit, an 
anomalous zone was identified between the 
Lower Freeport and Upper Freeport Coals. 



24 Observalions 



w-*- 




As this zone was in close proximity to the 
surface swale which defined the topo- 
graphic element of the photo-linear 
feature of interest, it was initially 
thought to be the subsurface expression of 
the photo-linear fracture trace. However, 
closer examination of the zone revealed 
several vein-like clastic infillings 
similar in character to the local under- 
clay and that the zone was offset to the 
east of the photo-linear feature. Other 
areas of the mine contained concentrations 
of these "clay vein" structures which were 
also not apparently related to fracture 
trace features. 

The three traverses of the siltstone 
unit also provided information on the 
fractures related to the mapped photo- 
linear feature as well as on the general 
distribution of joints in the area. In a 
total of 1,200 ft of highwall, 360 joints 
were identified. These features revealed 
a dominant orientation maximum between 320 
and 340 degrees (figure 5) which compares 
well with the orientation data provided by 
the previous analysis as well as other 
published joint surveys. Although there 
is variation between the three traverses, 
the orientation data are dominated by this 
distinct peak. It should be noted that 
the highwall was controlled by a face 
joint set which had a trend at 
approximately 90 degrees to the joint 
orientation maxima. The joint data is 
biased, therefore, because traverses along 
the highwall were much more likely to 
observe face perpendicular joints than 
face parallel joints. 

The joint data were also examined by 
preparing histograms of the number of 
joints per traverse interval (figure 6). 
This graphical technique revealed a peak 
to the east of the intersection point 
between the highwall and the mapped 
photo-linear feature. The number of 
joints per 20 ft traverse interval ranged 
from 2 to 1 1 with a mean of 6. This 
compares well to data for joint frequency 
in shales included in Nickelsen and Hough 
(1967). The peaks in figure 5 are not 
thought to be related to distinct fracture 
zones ; rather they appear to be related to 
stratigraphic features like the ones 
previously noted. 



360 Observalions 
10° Intervals 

Percent of Observation s 
10 




-»-E 



Figure 4 — Orientation rose of photo- 
linear fracture trace features. 



Figure 5 -- Orientation rose of all 
joint data. 



256 



10 



"Fracture Trace 
centerline 



5- 



TRAVERSE I (173) 



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TRAVERSE 2 (114) 



400 




600 



TRAVERSE 3 (73) 




200 400 

TRAVERSE DISTANCE (feet) 



200 



Figure 6 -- Traverse joint frequency 
distributions . 

Air-Injection Permeability Testing 
Because these air-injection tests make 
up the bulk of the quantitative data that 
was collected in this study, their analy- 
sis was critical to the objectives of this 
project. A total of 55 air-injection 
tests were conducted in six patterns of 
shotholes drilled into the siltstone 
unit. This particular aspect of the study 
gained importance with the discovery, by 
visual inspection of the highwall during 
the testing of Shothole Pattern 4 (see 
figure 1), that the photo-linear "fracture 
trace" was actually the surface expression 
of a stratigraphic swale in the Mahoning 
Coal. Although this discovery provided 
evidence that photo-linear features are 
related to geologic structures, this 
discounted even, further the "fracture 
trace" nature of the feature. 

The total range of permeability values 
determined in these tests is shown in 
figure 7. (1 foot squared (ft*) = 929 
centimeters squared (cm 2 )) These values 
are averages of each of the step values 
produced during the multiple-step ( abcba ) 
tests. The data appear to be log-normally 
and possibly bi-modally distributed. This 
variability is probably the result of 
several factors, the most important of 
which is the natural variations of the 
siltstone's permeability. A secondary 
factor affecting the data is the impact of 
mining activity (primarily blasting) on 



55 Observotions 
High- 12.419 xlCT'Mt* 
Low -0.530xl0""ft 2 
Mean-4.684xl0"ff2 



1.0 2.0 3.0 4.0 50 6.0 7.0 80 9.0 10.0 MO 12.0 

PERMEABILITY <k) x I0" 13 ft 2 

Figure 7 -- Distribution of air-injection 
permeability values. 

the in situ permeability values of the 
siltstone. This factor can be given a 
qualitative analysis by examination of a 
data subset, the permeability values 
obtained from Shothole Pattern 5 (Figure 
8). The distribution of values for the 
pattern reveals that permeability 
increases are highest along the highwall 
faces of the pattern. The data indicate 
that increases extend primarily from the 
recently shot area and that values higher 
than about 7 x 10" 13 ft 2 are probably 
significantly blasting affected. Analysis 
of other patterns supports these results 
and indicates that 2x + increases are 
likely 20 to 30 ft from the shot faces of 
the tested pattern. This relationship is 
probably dependent on the thickness of the 
shot overburden (O'Regan and Nguyen 1981). 

The multiple-step nature of the 
air-injection testing technique made it 
possible to examine other factors which 
could influence the permeability data. By 
preparing test step plots (see examples in 
figure 9 ) , departures from the ideal 
linear relationship between flow rate and 
pressure could be examined (Houlsby 
1976). Test 22 had one of the highest 
permeability values and steepest slopes 
recorded in this study. Of interest is 
the fact that the slope of the plot 
decreases with increasing pressure and 
flow. This slope change is also 

SAFETY 
/BERM\ |p 8 6 4 



/ 
/ 
/ 
/ 
/ 
_ / 



H / 

d / 
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x / 

/ 
/ 
/ 
/ 
/ 



Tested Shothole • 
Untitled Sholhole 



All Values From 
Constant -Head, 
Air -Injection Tuts 




All Permeability Value 
Are k s 10 "It* 



SAFETY BERM 



HOLE MUM8ER 1-6 



V/ /////////////////////////////// 

HIGHWALL (OLD) 

Figure 8 -- Isopermeability map of 
Shothole Pattern 5 . 



257 



Test 22lobo) 
Test 23(obcba) 
Test 34(abcba) 
Test 44(abcba) 



k-l2.209xl0" 13 ftz 
k- 7.722 x ICris ft* 
k«4.2l9 xlO-'Mt 2 
k-2.129 xlO" 13 ft 2 



Test 54(obcdcba) k- 0.557 xlO" 13 ft 2 
(with transducer dota — o — ) 



direction of pressure step 
34 




20 
PRESSURE (psi) 

Figure 9 -- Step plots of selected air- 
injection tests. 

illustrated by other high permeability 
tests, such as Test 23, and it is 
attributed to either turbulent flow 
conditions developing in the fractured 
porous media or to an improved packer seal 
in the borehole. Test 34, which has a 
moderate slope and permeability value, 
illustrates the idealized Darcian 
relationship between pressure and flow 
rate. As pressure increases in the test 
interval, the flow of injected fluid from 
the test interval increases in an amount 
which is dependent on the permeability of 
the material. Test 44 records a slope 
increase with increasing pressure and flow 
for a relatively low permeability test. 
This change is probably the result of the 
widening of a fracture in the siltstone or 
the leakage of fluids past the packer 
unit. Test 54 recorded a very low 
permeability value and step plot slope. 
This test was designed to test uphole 
pressure gauging equipment with downhole 
transducer readings. The test was 
conducted in a wet shothole, and the 
gradual increase in the observed perme- 
ability values with increasing pressure 
and testing time is thought to be due to 
the displacement of water from pores and 
fractures which resulted in increasing gas 
transmission from the test interval. 
Further analysis to determine the degree 
of influence of these test factors was not 
possible with the available testing 
technique . 

Although it should have been possible 
to apply a Klinkenberg gas slippage 
correction term to the multiple-step test 
data, it proved to be impractical due to 
non-linear variations in the permeability 
values. The gas slippage correction was 



developed by Klinkenberg (1941) to account 
for the increasing interference, and 
related decrease in observed permeability, 
between gas molecules as increasing 
injection pressure makes them behave as an 
incompressible liquid. In this situation, 
because the air was not sufficiently 
compressed to behave as a liquid in these 
tests, it is possible that the measured 
permeability values are affected by 
Klinkenberg slippage. 

Other Hydrologic Tests 

Because of the restricted number of 
monitoring boreholes available to be 
tested by the falling head and constant 
head water-based tests, as well as field 
conditions which made some of the holes 
unsuitable for testing, only two hydraulic 
conductivity values could be obtained for 
the siltstone unit. However, the close 
correlation between these values from Hole 
2 (3.9 and 5.1 x 10-7 ft/sec for a 
falling-head and constant-head test 
respectively) indicates that methods are 
comparable in this environment. These 
data and methods may also be comparable to 
the air-injection technique. This is 
supported through the use of a data table 
from Freeze and Cherry (1979) that allows 
all the air- and water-based test values 
to be compared. This table (figure 10) 
shows that the permeability and hydraulic 
conductivity values are limited in range 
and appear to represent the equivalent of 
a permeable sandstone. This does not seem 
unreasonable because the fractured silt- 
stone unit may be hydrologically similar 
to such a lithology. 

Additional uses of the air-injection 
technique provided some interesting 
results. Falling-head air-injection tests 



Rocks 



Unconsolidated 
deposits k 



2 o 

J5| 

e i 

n o c i 

^ c _ r 



■s'e-i 

■o o o : 

»£••! 

»J«oo 
u. | Jtj-o 



tfl 



V) 

If 



l! 



O ,_ u, 

O O 3 

sis 



\ 



(ft 2 ) (cm 2 ) 
io _s rio~ 3 

I0" 4 

1-I0" 5 

lO" 6 

I0" 7 

icr 8 

10-9 

io-'° 

lO" 12 
I0"' 3 
IQ-14 



-|0" 7 

■I0" 8 

10-9 

10-10 

10"" 
I0" 12 
lO" 13 
lO" 14 
lO" 15 
lO"' 8 

ho- 17 



i -i8 -icr 1 ' 



L| -I9 



K K K 

(cm/s)(ft/s) (gal/day/ft 2 ) 
rio 2 

r-10 6 



10 

I 

lO' 1 

I0" 2 

I0" 3 

|0"4 

ho- 5 



I0' 6 "lO" 8 



lO' 8 

io-'o 

IQ-16 HO-" 



io-i 

I0" 2 
lO" 3 
10*4 
10-5 
-I0" 6 
I0" 7 



| -8 -IQ-IO 



lO"" 
lO"' 2 
10-13 



Range of Air-Injection 
Test Values 



I0 5 

10* 

I0 3 
I0 2 
10 

-I 

-10"' 
I0" 2 
I0" 3 

ho- 4 
io-= 

I0' 6 
I0" 7 



x = Water Test Values 



Figure 10 — Range of test values for 
permeability (k) and hydraulic 
conductivity (K) (modified from Freeze 
and Cherry 1979) 



258 



(Blankenship and Stickney 1983) appeared 
to underestimate permeability values by 
two orders of magnitude. Air-injection 
interference testing conducted in Holes 1 
and 2 (see figure 1) was more successful. 
During injection of air through a borehole 
packer inserted in Hole 2 while monitoring 
water levels about 20 ft away in Hole 1 , 
it was observed that a water level 
fluctuation occurred in less than 10 
minutes in the monitored hole. This was 
surprising as it had taken nearly 90 
minutes for water injected in Hole 2 (at 
nearly the same pressure as the air) 
during the constant-head water-injection 
test, to cause even a minor change in the 
water level of Hole 1 . The air-injection 
test also caused a much more substantial 
change in the observed water levels, 
nearly 1.7 ft as opposed to a few tenths 
of a foot. While the significance of 
these results is unclear, it is clear that 
injected air is capable of markedly 
affecting a ground-water flow field under 
some circumstances. 

CONCLUSIONS AND RECOMMENDATIONS 

The investigation of fracture features 
and in situ permeability values conducted 
at the Browncrest III surface coal mine 
succeeded in addressing a number of 
research objectives. The conclusions and 
recommendations of this work may assist 
future studies. 

Conclusions 

1. The photo-linear "fracture trace" 
orientation peak, between 320 and 330 
degrees, compares well to the orientation 
of local tributary streams and the maxima 
of other fracture trace studies for this 
area . 

2. The joints mapped in this study had 
an orientation maxima between 320 and 340 
degrees. This agrees well to the fracture 
trace orientation peak, as well as to the 
orientation of the local stream valleys 
and to other published joint data. As a 
result, it is believed that all the 
features may be related to a common 
structural origin. 

3. Measurements of joint frequency per 
highwall traverse interval revealed that 
joint concentrations appear to be related 
to specific stratigraphic features rather 
than fracture traces. 

4. The photo-linear feature that was 
to be the focus of this study was found to 
be unrelated to a zone of fracture 
concentration. This feature was found to 
be related to a stratigraphic swale, 
possibly an old distributary channel, in 
the Mahoning Coal . 

5. The air-injection technique was 
found to be quite consistent, durable, and 
mobile in operation. It was capable of 
conducting tests for a specific range of 
conditions: a) boreholes at depths of up 
to 70 ft, b) hole diameters of 4 to 7 
inches, and c) injection pressures and 
flow rates well in excess of test 
parameters . 

6. The in situ permeability values 
observed in this study ranged from 0.5 to 
12.5 x 10" 13 ft 2 and appeared to be 



log-normally and possibly bi-modally 
distributed. Mining activity was 
determined to affect the permeability 
values, increasing values by 2x + up to 30 
ft from mine highwalls. As a result, the 
background permeability of the siltstone 
unit is probably in the range 1.0 to 7.0 x 
10- i 3 ft* . 

7. Factors affecting the air tests 
were examined through graphical analysis 
of the test data. Turbulent flow condi- 
tions are thought to have the greatest 
impact on the values, although partially 
saturated conditions (water in the 
siltstone) and Klinkenberg slippage may 
also significantly affect the values. 

8. Hydraulic conductivity values 
obtained by falling-head and constant-head 
water-based tests from the siltstone unit 
compare well to the air-based permeability 
values. These values indicate that the 
fractured siltstone unit has a character 
similar to a moderately permeable 
sandstone . 

9. Additional uses and modifications 
of the air-injection technique were also 
investigated. Interference testing offers 
interesting data, although it is not 
possible to evaluate this information at 
this time. 

Recommendations 
Although this study was not able to 
quantify the subsurface nature of a 
photo-linear fracture trace, the methods 
presented here define a means of achieving 
this objective. However, the reliability, 
speed, safety, precision, and economics of 
these techniques may require further 
investigation. The photo-linear analysis 
may identify geologic structures, as well 
as nongeologic ones, that are not related 
to zones of fracture concentration in the 
subsurface. The highwall surveys are time 
consuming, occasionally dangerous, and are 
complicated by blast fractures and the 
relationship of the fractures to the 
highwall orientation. Many factors affect 
the air-injection technique. However, 
most of them are inherent to the technique 
and can only be corrected with improved 
analysis routines or increasingly costly 
instrumentation. Therefore, although this 
technique may be appropriate under 
circumstances where a high density of data 
is required, it may not be appropriate for 
more general investigations (Shuman 1987). 

ACKNOWLEDGEMENTS 

This project was made possible, in 
part, by grants-in-aid from Sigma Xi , and 
Chevron's Field Research Fund. Other 
portions of this study were supported by 
The Pennsylvania State University's 
Krynine Memorial Fund, Mineral 
Conservation Section, and Coal Research 
Section. The assistance of numerous 
colleagues during this project is also 
gratefully noted. 



259 



LITERATURE CITED 

Blankenship, D.A. and R.G. Stickney, 1983 
Nitrogen gas permeability tests at 
Avery Island Report for Battelle Proj 
Management Div. Office of Nuclear 
Waste Isolation, ONWI-190, 22pp. 

Cooper, H.H., J.D. Bredehoeft and I.S. 
Papadopulos, 1967. Response of a 
finite-diameter well to an instant- 
aneous charge of water Water Res. 
Research, vol. 3, no. 1, pp. 263-269. 

Fetter, C.W., Jr., 1980. Applied 

Hydrogeology Charles E. Merrill 
Publishing Co., Columbus, OH, 488pp. 

Freeze, R.A. and J. A. Cherry, 1979. 
Groundwater Prentice-Hall, Inc., 
Englewood Cliffs, N.J., 604pp. 



Parizek, R.R., 1976. On the nature and 
significance of fracture traces and 
lineaments in carbonate and other 
terranes in: Karst Hydrology and Water 
Resources, Proceedings of the U.S. - 
Yugoslavian Symposium, Dubrovnik, 
Water Resources Publications, Fort 
Collins, CO, pp. 41-100. 

Pearson, R. and M.S. Money, 1977. 

Improvements in the Lugeon or packer 
permeability test Quarterly Journal of 
Eng. Geology, vol. 10, pp. 221-239. 

Price, M. , B. Morris and A. Robertson, 
1982. A study of intergranular and 
fissure permeability in Chalk and 
Permian aquifers, using double-packer 
injection testing Journal of 
Hydrology, vol. 54, no. 4, pp. 
401-424. 



Houlsby, A.C., 1976. Routine interpre- 
tation of the Lugeon Water-Test 
Quarterly Journal of Engineering 
Geology, vol. 19, pp. 303-313. 

Klinkenberg, L.J., 1941. The permeability 
of porous media to liquids and gases 
in: Drilling and Production Practice, 
American Petro. Inst., pp. 200-213. 

Lattman, L.H., 1958. Technique of mapping 
geologic fracture traces and 
lineaments on aerial photographs 
Photogrammetric Engineering, vol. 19, 
no. 4, pp. 568-576. 

Lattman, L.H. and R.H. Matzke, 1961. 
Geologic significance of fracture 
traces Photogrammetric Engineering, 
vol. 27, pp. 435-438. 

Lattman L.H. and R.P. Nickelsen, 1958. 
Photogeologic fracture-trace 
mapping in the Appalachian Plateau 
American Assoc, of Petro. Geologists 
Bulletin, vol. 42, pp. 2238-2245. 

Meiser, E. and T. Earl, 1982. Uses of 
Fracture Traces in Water Well 
Location: A Handbook Office of Water 
Research and Technology, U.S. 
Department of the Interior, 55pp. 

Nickelsen, R.P. and V.N. D. Hough, 1967. 
Jointing in the Appalachian Plateau 
of Pennsylvania The Geological Soc. of 
America Bull., vol. 78, pp. 609-630. 

O'Regan, G.J. and V.V. Nguyen, 1981. A 
study of highwall damage caused by 
overburden blasting in a strip coal 
mine Australasian Institute of Mining 
and Metallurgy, Central Queensland 
Branch, Symposium on Strip Mining - 45 
m and Beyond, pp. 191-200 



Shuman, C.A., 1987. Fracture studies and 
in situ permeability testing with 
borehole packers, Unpublished M.S. 
Thesis, The Pennsylvania State 
University, University Park, PA, 185p. 



260 



MODELING SULFATE RETENTION IN A LAKE RECEIVING ACID MINE DRAINAGE 1 



Alan T. Herlihy, Aaron L. Mills, and Winston Lung' 



Abstract. -- Bacterial sulfate reduction in the 
sediments of Lake Anna, VA combined with 
geochemical processes in the water column remove 
half the S0 4 entering the lake in acid mine 
drainage (AMD) from Contrary Creek. In addition to 
S0 4 removal, the pH of the water is increased 
from about 3.5 to 6.0 within the boundaries of the 
contaminated arm. A simple water quality model was 
constructed to predict the spatial and temporal 
distribution of S0 4 in the lake water and to help 
identify the important processes controlling the 
distribution of the pollution in the lake. The 
model successfully predicted distribution of S0 4 
in most locations, except for those closest to the 
mouth of the AMD stream. The model demonstrated 
that chemical stratification of the lake water was 
more important than thermal stratification near the 
mouth of the acid mine stream and that a strong 
chemical gradient there inhibits vertical mixing. 
Maintenance of the AMD plume near the sediment 
surface where the biological activity occurs likely 
enhances S0 4 removal. The proportion of S0 4 
retained in the Contrary Creek arm (S0 4 retained 
/ S0 4 influx) was 0.48, consistent with other 
lakes which actively retain sulfur in the 
sediments. The comparatively short residence time 
of this arm of Lake Anna (ca. 100 days) yields an 
S0 4 -removal coefficient of 12 to 14, which is 
over an order of magnitude higher than reported for 
lakes acidified experimentally or by precipitation. 
In those lakes diffusion is assumed to be the major 
mechanism of S0 4 transport to the sediments where 
most SR occurs. In Lake Anna, the model results 
demonstrated that some other mechanism plays an 
important role in transporting the AMD constituents 
from the lake water to the sediments. The amount 
of AMD neutralized by the biogeochemical processes 
in this lake suggests that some impoundments might 
be appropriate for the renovation of AMD- 
contaminated waters. 



261 



INTRODUCTION 

In Lake Anna, VA, successful 
neutralization of AMD from a series of 
abandoned pyrite mines occurs within one 
arm of the lake within 2 km of the point 
where Contrary Creek enters the lake. 
Contrary Creek has an annual average pH 
about equal to 3.2, S0 4 2 about equal to 1 
to 20 pmol/L, and total iron about equal 
to 10 to 50 mg/L. On average, 48 % of the 
S0 4 that enters the lake from Contrary 
Creek is removed in the first two 
kilometers of the lake (Herlihy et al., 
1987). Concomitantly, the pH rises to 
approximately 6 and the iron levels drop 
to levels similar to uncontaminated arms 
of the lake. This homeostatic renovation 
of the water has been attributed to 
anaerobic bacterial activities, 
specifically sulfate reduction (SR) in the 
sediments underlying the contaminated arm 
of the lake (Mills 1985, Mills and Herlihy 
1985, Herlihy and Mills 1985, Herlihy et 
al. 1987; Mills et al. in press). 

The establishment of anaerobic 
conditions, SR, and the resultant 
precipitation of metal sulfides are 
significant in increasing the pH and 
reducing the AMD-derived iron and S0 4 
concentrations in the lake water. 
Alkalinity generation from SR occurs 
according to the equation: 



2CH 2 + S0 4 ' 



H 2 S + 2HC0 3 



2- 



(1) 



Evaluation of the relative amount of 
neutralization that SR can provide is 
related to the amount of S0 4 removed 
from the water. Although SR plus dilution 
provides a thorough cleansing of water in 
the Contrary Creek Arm of Lake Anna, 
prediction of the efficiency of this 
homeostatic process in other systems must 
rely on an adequate quantitative model to 
test other systems prior to construction 
of new impoundments or contamination of 
pre-existing waters. The present study 
applied a modified version of the WASP 
model (Water Quality Analysis Simulation 
Program) to simulate the conditions in 
Lake Anna, with the intent of applying the 
model to other acidified impoundments in 
the future. 



Paper presented at the 1988 Mine 
Drainage and Surface Mine reclamation 
Conference sponsored by the American 
Society for Surface Mining and Reclamation 
and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement) , April 
17-22, 1988, Pittsburgh, PA. 

2 Alan T. Herlihy is Research Scientist 
with USEPA, Corvallis, OR, and Aaron L. 
Mills and Winston Lung are professors of 
Environmental Sciences and Civil 
Engineering, respectively, University of 
Virginia, Charlottesville. VA. 



METHODS 
Model Description 

WASP is a general model capable of 
handling one, two, or three dimensional 
time-variable calculations (DiToro et al. 
1983). The water body to be modeled is 
divided into segments and mass balance 
equations for each segment are constructed 
for the constituent of interest. The mass 
balance equations are solved using a 
finite difference technique with a 
backwards-dif f erence approximation in the 
spatial plane and a forward-difference 
approximation in the temporal plane. The 
finite difference equations are integrated 
using a second-order Runge-Kutta method. 
WASP has been used successfully in 
modeling lake acidification in Bickford 
Reservoir, and Woods and Panther Lakes 
(Lung, 1987) 



Application of WASP to Lake Anna 

Only the area of the lake between the 
mouths of Freshwater and Contrary Creeks 
and the Route 6 52 bridge near station A2 
was included in the model (fig. 1). To 
implement WASP in the Contrary Creek arm 
of Lake Anna, this part of the lake was 
divided into eight segments (fig. 1). 
Epilimnion and hypolimnion segments were 
included for areas of the lake around 
stations C2 , C5 , and A2 . The Freshwater 
Creek section of the lake was treated as 
one segment as was the small segment of 
lake near the mouth of Contrary Creek 
around station CI. Sulfate concentration 
was the only constituent modeled, as S0 4 
retention approximates the neutralization 
process for the lake. 

The flows into each segment of the 
model are shown diagrammatically in figure 
1, and the mass balance and discharge and 
loading equations are given in table 1. 
Sulfate from Contrary Creek enters into 
segment 1, and 9 0% flows into segment 5 
(hypolimnion of station C2) with the 
remainder flowing into segment 2 
(epilimnion of station C2). Water from 
segment 2 flows into segment 3 , and then 
segment 4 along the surface of the lake 
(epilimnion segments of stations C5 and 
A2) before exiting out the outflow. Water 
in segment 5 flows into segment 6 and then 
segment 7 along the bottom of the lake 
(hypolimnion segments of stations C5 and 
A2) before exiting out the outflow. 
Sulfate is mixed between the epilimnion 
and hypolimnion of each station (segments 
2 and 5, 3 and 6, and 4 and 7) by vertical 
eddy diffusion. Flow percentages were 
calculated from lake cross sectional areas 
and current velocity observations. Water 
and S0 4 from Freshwater Creek enter into 
segment 8, and then 3 8% flows into segment 
4 and the remaining 62% into segment 7. 
Each surface segment (1,2,3,4 and 8) 
received S0 4 and water from 



262 



Figure 1. — A. Map of the Contrary Creek 
arm of Lake Anna. Water flows in an 
easterly direction. The area shown 
represents about 13% of the total surface 
area of the impoundment which lies to the 
northeast and southeast of the arm shown. 
B. Schematic diagram of flow routing and 
segment location for the model applied to 
the Contrary Creek arm of Lake Anna. 



Conlrary 
Creek 




CI 



C2 



CS 



A2 




OUTFLOW 



DEPTH (■) 



FRESHWATER 
CREEK 



1.0 

— I— 



l.S 



2.0 



2.5 



DISTANCE DOWNSTREAM FROM MOUTH OP CONTRARY CUB (ks) 



precipitation and direct input (overland 
flow and ephemeral streams) and lost water 
due to evaporation. Sulfate is removed 
from the segments overlying sediments 
(1,5,6,7 and 8) by SR minus sulfide 
oxidation (Herlihy et al. 1987). 

The data needed by WASP to run the 
model (described later) included the 
volume and initial S0 4 concentration of 
each segment, the waterflows between each 
segment, the eddy diffusion between 
segments, and the external S0 4 z load into 
each segment of the model. The kinetic 
term for SR and the associated 
coefficients were also input. Finally 
boundary conditions indicated the sulfate 
concentration of the outflow from segments 
4 and 7. 



Flow, Load and Concentration Data 

The model was calibrated using data 
collected during the 1984 water year 
(October 1, 1983 to September 30, 1984) 
and verified using data collected during 
the 1983 water year. Discharges and loads 
were input into the model at 5-day 
intervals using the mean value of the five 
daily discharges and loads. Discharge and 
load data were taken from the measured 
sulfate budget for the lake (Herlihy et 
al . 1987), recalculated for each segment. 
Initial conditions and boundary conditions 
were taken from the observed lake water 
sulfate concentrations. Segment volumes 
and surface areas were obtained from 
Bruckner (1986) . 



263 



Table 1. -- Mass balance and loading and discharge equations 
used in the application of the modified WASP model to Lake Anna. 



MODEL MASS BALANCE EQUATIONS 



V 1 dC 1 /dt = W 1 -Q 12 C 1 -Q 15 C 1 -k 8ulf C 1 A 1 
V 2 dC 2 /dt = W 2 ♦ Q U C, - Q 23 C 2 + K s A 26 /L 25 (C 5 -C 2 ) 

W dt = W 3 + «23 C 2 " ^S4 C S + K , A 36/ L 36( C 6- C 3' 
V 4 dC 4 /dt = W 4 + Q S4 C 3 + Q 84 C g - Q 40 C 4 + K i A 47 /L 47 (C 7 -C 4 ) 



«56 = <> 



16 



%8 = Q fc + S 8«fc + A 8 Q prec " A 8 Q evap " A 8 Q store 



V c dC K /dt = C. C, - Q SR C R + K A, R /L,-(C,-C I .) - k, ,..C R A 



•15 1 



i 25' 2S V 2 



sulf 5 25 



V R dC fi /dt = Q sfi C R - Q fi7 C R + KAJIJC.-C.) - k , ,..C A 



I 36' 36 v 3 



sulf 6 36 



V 7 dC 7 /dt = Q 67 C 6 + Q 87 C 8 - Q 70 C ? + Ki A 47 /L 47 (C 4 -C 7 )- 



sulf 



'7*47 



Note that segment represents external boundary condition. 



V. = Volume of segment i (m ). 

C. = Concentration of sulfate in segment i (mol/m). 

Q.. = discharge from segment i to segment j (m /day). 

W. = external sulfate loading into segment i (mol/day). 

K = vertical eddy diffusion coefficient (m /day). 

A.. = surface area between segments i and j (m ). 

A. = surface area of segment i (m ). 

L. . = sum of the depths of segments i and j (m). 



'J 



discharge from segment i to segment j. 



Q = discharge from Contrary Creek. 

Q_ = discharge from Freshwater Creek. 

Q = water input from precipitation. 

Q = water loss due to evaporation. 

evap 
Q = discharge due to change in lake storage volume. 

S. = ratio of direct input watershed area in segment i 
to the Freshwater Creek watershed area. 

A. = ratio of lake surface area in segment i to total 
lake surface area. 



DISCHARGE EQUATIONS 



LOADING EQUATIONS 



Qn, = Q +S.Q, +A.Q -A,Q - A.Q . 

01 ^cc 1 Mc l^prec l^evap 1 store 



W, = W + A.W + S,W, 

1 cc 1 prec 1 fc 



^02 = S 2<»fe + A 2%rec _ A 2 Q evap " Mstore 
%3 = S 3«fc + A 3 Qprec " ^evap " A 3 Qstore 



W, = A.W + S.W, 

2 2 prec 2 fc 

W. = A-W + S.W. 

3 3 prec 3 fc 



04 4 fc 4 prec 4 evap 4 store 



W . = A.W + S.W, 
4 4 prec 4 fc 



Q 15 = 0.98 • Q 01 



W„ = W, + A„W + S„W, 

8 fc 8 prec 8 fc 

W. = external sulfate loading into segment i. 

W = sulfate loading from Contrary Creek. 

cc 

W, = sulfate loading from Freshwater Creek. 

W =sulfate loading from precipitation. 



Vertical Eddy Diffusion Coefficients 

Sulfate transport via vertical eddy 
diffusion was calculated as the product of 
the vertical diffusion coefficient (K ) 
and the interfacial area between the 
epilimnion and hypolimnion divided by the 
average depth of the two layers. There 
are no data about the magnitude of the 
vertical eddy diffusion coefficient in 
Lake Anna, so a value of 0.05 cm /sec was 
chosen as a first approximation of K z , 
based on a range of literature values for 
a number of similar and dissimilar lakes. 



where C^ is the 
segment i (mols 
area of segment 
of segment i (m ) 
removal coeffic 
removal only ta 
overlying sedime 
k sulf describes 
sulfate reduction 
The coefficient 
piston velocity 
of water per year 
sediments to ac 
sulfate removal. 



S0 4 
/m 3 ) , 
i (m 2 

and 
ient 
kes 
nts. 
the 

minu 

can 
relat 

must 
count 



2- 



concentration in 
A^ is the surtace 
) , V^ is the volume 

^sult * s fc ^ e su lf ate 
(m/yr) . Sulfate 

place in segments 
As modeled here, 
net reaction of 

s sulfide oxidation. 

be thought of as a 

ing how many meters 

be processed by the 

for the observed 



Sulfate Removal Kinetics 

Sulfate removal was modeled by the 
equation 



Baker et al . (1986) presented an 

equation for calculating k su jf based on 

the mean depth, water residence time, and 
sulfate retention of a lake: 



v i * <3 Ci /dt= k su if 



c i * A i 



(2) 



'SUlf 



R * z / (t w * (100-R) ) 



(3) 



264 



In ,this equation, R is the retention of 
as percentage of input, t w is the 



SO, 



water residence time (yr.), and z is the 
mean depth (m) . Values of k SU ]f in the 
present study were then calculated using 
the budget data from Herlihy et al. 
(1987). The average depth was calculated 
by dividing the Contrary Creek arm lake 
volume by its surface area (4.4 x 10° m / 
1.1 x 10* m 2 = 4 m) . Calculated values of 
^sulf were 12 - 7 m/year in the 1983 water 
year and 14.1 m/year for the 1984 water 
year. These coefficients were used in the 
model in the kinetic expression for 
sulfate removal as shown in equation 2. 



RESULTS 



Model Calibration 

Average epilimnetic and hypolimnetic 
sulfate concentrations predicted by the 
model for the 1984 water year were in good 
agreement with observed data. However, 
the model predicted little chemical 
stratification between epilimnion and 
hypolimnion usinga K of 0.05 cm /sec. 
The observed S0 4 data showed a strong 
stratification at station C2 during most 
of the year and at station A2 in the 
winter. Therefore the model was rerun 
after changing the spatial and temporal 
values of K z so that a reasonable fit to 
the data set was obtained. The K z values 
for stations C5 and A2 were lowered to 
0.01 cm /sec. In order to obtain a good 
fit to the observed station C2 data, the 
K z had to be lowered even more to 0.002 
cm /sec. It was also necessary to change 
the flow routing so that 98% of the water 
from Contrary Creek coming from segment 1 
went into segment 5 (C2 hypolimnion) . In 
order to fit the observed winter sulfate 
distribution at station A2 , it was 
necessary to decrease all three K z s to 1 x 
10 cm /sec from December 2 5 to February 
15 to reflect the reduction in vertical 
diffusion due to ice cover on the lake. 
The ice cover would lower K z by stopping 
any wind mixing. 

To make the model more realistic, 
^sulf was set t0 ° during tne months of 
January and, February to reflect the 
cessation of sulfate removal as observed 
previously (Herlihy, 1987). During 
October, November, and December, k su ^f was 
decreased linearly with time from the 
maximum summer value to zero. Similarly 
during March, April, and May, ^sulf was 
increased linearly with time from zero to 
the maximum summer value. The maximum 
summer k S(J if (in effect from June through 
September) was calculated so that the 
value calculated in equation 3 was the 
annual average k SU 2f. 



were in fairly good agreement with the 
observed data (fig. 2). The model failed 
to predict the low hypolimnetic S0 4 
concentrations in June and July. The model 
also underestimated S0 4 concentration in 
the epilimnion from October to April and 
overestimated it from July to September. 



2000- 








C2 




OS9ERVCD EPILIMNION 
PREDICTEO EPILIMNION 

• OMERVED HYPOLIMNION 
PREDICTED HYPOLIMNION 


1SOO- 














a 










• 


* • 


lOOO- 






V • 




• 




o' 

CO 500 
l_l 


Jt 


o , 





» °o 


o 
o 


O- N_ - 

0- o • 

,xr • ,8 




0- 


_ 


1 


i ... i ,. 


„ i ... i„.. 1 




1 ...„ 1 .... 1 1 1 



1983-1984 

Figure 2. — Observed and model-predicted 
sulfate concentrations (nmol/L) in 
the epilimnion and hypolimnion at 
station C2 (segments 2 and 5) during 
the 1984 water year. 



The model 
concentrations a 
better than it di 
the observed SO 
within 10-20% 
concentrations, 
predicted th 
stratification i 
(fig. 4) . From 
the predicted S0 4 
hypolimnion were 
observed concentr 



predicted 
t station C5 

d at station C2 

2 
4 



SO. 



2- 



(fig. 3) 

Most of 

concentrations were 

of the predicted 

At station A2 the model 

e observed S0 4 

n January and February 

June through September, 



concentrations in the 
about 50% lower than the 
ations. 



C5 



2000 



1000- 



o 

o> 500- 



ORSERVED EPILIMNION 
PREDICTED EPILIMNION 
ORSERVED HYPOLIMNION 
PREDICTED HYPOLIMNION 




OCT ' NOV ' DEC ' JAN ' FED ' MAR ' APR ' HAT JUH JUL AUO SEP 

1983-1984 

Figure 3. — Observed and model-predicted 
sulfate concentrations (nmol/L) in 
the epilimnion and hypolimnion at 
station C5 (segments 3 and 6) during 
the 1984 water year. 



Sensitivity Analysis 



After 
predicted 



these 



SO 



^- e 



calculations, the 
concentrations in both 



epilimnion and hypolimnion at station C2 



To understand the sensitivity of the 
model to k slJ if and K z a sensitivity 
analysis of ' these parameters was 



265 



OH1RV10 EPILIMNION 
PB(OICTEO EPILIMNION 
ONSIRVEO HYPOLIMNION 
PREDICTED MvrOUMMIO" 




2000 



1500- 



O 

(0 500 



C2 Eplllmnlon 

VARYING K, (oa^/MC) 







OCT NOV DU JAN ' Fit ' tMII ' APR ' HAT ' JUM 

1983-1984 



Figure 4. -- Observed and model-predicted 
sulfate concentrations ((imol/L) in 
the epilimnion and hypolimnion at 
station A2 (segments 4 and 7) during 
the 1984 water year. 



performed. The predicted sulfate 
concentration in the hypolimnion at 
station A2 was very sensitive to small 
changes in k gu if (fig. 5). The solid line 
in the middle of figure 5 was the 1984 
value of k su if calculated from equation 3 
(14.1 m/yr). The line predicting higher 



SO 



annu 



a* 



concentrations represents an 
^sulf °^ i- 85 m/yr and the lowei 
concentration line had a value of 7(5.0 
m/yr. All of the k S nif values in this 
exercise were set to during the winter 
months and varied with time as described 
above. The sensitivity analysis shows 
that the range of acceptable values of 
^sulf * s sma H' about 5-20 m/yr. 



800T 



600- 



A2 Hypolimnion 

V»BTINQ «, u ,| In/Vfl 




OCT NOV ' DEC JU ' FEI ' Si ' APR ' MAT JUN 

1983-1984 



JUL AUG SEP 



Figure 5. -- Model-predicted sulfate 
concentrations in the hypolimnion of 
station A2 (segment 7) at varying 
values (m/ yr) of the sulfate removal 
coefficient (k SU 2f). 

The sensitivity analysis for K z showed 
that large (order of magnitude) changes 
made small, but significant changes in the 
predicted epilimnetic S0 4 concentration 
at station C2 (fig. 6). With a K z of 1 x 
10 4 cm /sec there was little or no mixing 
between epilimnion and hypolimnion. Thus, 
predicted S0 4 concentrations were low 
and remained fairly constant with time 



Figure 6. — Model-predicted sulfate 
concentrations (nmol/L) in the epi- 
limnion of station C2 (segment 2) at 
varying values (cm / sec) of the 
vertical eddy diffusion coefficient 
<K Z ). 

because waterflow through the epilimnion 
was low (only 2% of Contrary Creek 



inflow) , and there was little SO, 



influx 



from 
There 



direct input and precipitation, 



was more SO, 



transfer between 



hypolimnion and epilimnion when K z was 1 x 
10~ 2 cm /sec. The effects of different K z 
values on the predicted hypolimnetic S0 4 
concentrations at station C2 were less 
than 10% (data not shown). 

Model Verification 



The model was ve 
collected during the 1 
the same values of 
calibration. The valu 
the verification was 12 
varied temporally as de 
predicted S0 4 conce 
hypolimnion at station 
the observed concentrat 
were similar (fig. 
epilimnetic S0 4 co 
scattered around 
concentration line. 



rified using data 

9 83 water year and 

K z used in the 

e of k SU 2f used in 

.7 m/yr, and it was 

scribed above. The 

ntrations in the 

C2 were higher than 

ions but the trends 

7) . Observed 

ncentrations were 

the predicted 



OBSERVED EPILIMNION 
PNIOICTEO EPILIMNION 
OMERVEO HYPOLIMNION 
PNIOICTEO HYPOLIMNION 




OCT ' NOV ' OCC ' JAN ' Pli ' MA* ' APR ' MAT ' JUN ' JUL ' AUO ! SEP ' 

1982-1983 

Figure 7. — Observed and model-predicted 
sulfate concentrations ((jmol/L) in 
the epilimnion and hypolimnion at 
station C2 (segments 2 and 5) during 
the 1983 water year. 



266 



Except for the late winter and early 
spring months, the observed and predicted 
S0 4 concentrations at station C5 were in 
good agreement (fig. 8). The high S0 4 
concentrations predicted in the 
hypolimnion in the winter were not seen in 
the observed February and ..arch data. The 



predicted 



and 



observed 



SO 



4 



concentrations at station A2 were also 
similar except for the predicted 
hypolimnetic sulfate peak in February and 
March that was not present in the observed 
data (fig. 9). Model efficiency was 
calculated using the verification data 
(predicted and observed) from March 
through September 1983. F values (sum of 
the squares of the observed data about the 
mean divided by the sum of the squares of 
the predicted data-observed data) were 
very low (ranging from 0.19 in the A2 
hypolimnion to 2.03 in the A2 epilimnion); 
there was as much variance in the model 
prediction about the observed data as in 
the observed data about the mean. 



2000-1 



cs 



O OIMRVEO CmUMNIM 

'«f OICTEO EPILIMNION 

• OMMVCD HfPOllHNION 
PREDICTED HTPOUMNIOM 




Figure 8. — Observed and model-predicted 
sulfate concentrations (|imol/L) in 
the epilimnion and hypolimnion at 
station C5 (segments 3 and 6) during 
the 1983 water year. 



oattflvcD epilimnion 

PNEDICTEO EPILIHMIOM 
OMEP.VEO HTPOUMNION 
PREDICTED HTPOLIMNIOM 




OCT NOV OIC 



JAM FH ' MAM ' iPl ' MAY JOM 

1982-1983 



Figure 9. — Observed and model-predicted 
sulfate concentrations (nmol/L) in 
the epilimnion and hypolimnion at 
station A2 (segments 4 and 7) during 
the 1983 water year. 



DISCUSSION 

In order to fit the model to the 
observed data, it was necessary to assume 
that 98% of the AMD inflow went into the 
hypolimnion of C2 and that K„ at station 
C2 was lower than the other stations. At 
C2 the more AMD impacted water in the 
hypolimnion has a higher density than the 
epilimnetic water so there is a barrier to 
mixing that would cause a lower K z . At 
stations C5 and A2 the chemical gradient 
has been reduced by S0 4 and metal 
removal, dilution, and mixing so that the 
chemical barrier to mixing is reduced and 
a higher K z would be expected. At 
certain times in the year, especially at 
station C2 , the observed and predicted 
S0 4 concentrations were not very close. 
It is likely that the actual K z varies 
more temporally than the K z used in the 
model. Factors such as storm events, and 
large changes in air temperature, and wind 
direction could cause large changes in K„ 
for short periods of time. A storm event 
would carry in a large amount ot dilute 
water destroying a chemical stratification 
and perhaps causing turbulent mixing. 
Similarly, the S0 4 maximum predicted by 
the model in January and February 1983 was 
not seen in the observed data. The ice 
cover in 1983 was much less than it was in 
1984. if ice cover was the factor 
reducing K„ in the winter of 1984, it is 
likely that the K in the winter of 1983 
should not have been as low as it was 
modeled. Actual measurements of K„ with 
time and the percent of AMD inflow 
entering the hypolimnion would enhance the 
predictive ability of the model. 



Baker et al . (1986) reported a 
k i f of 0.46 (+ 0.30) m/yr for 
softwater lakes. They found that k su -, 
inversely related to the lake's residence 
time. Lakes with a long residence time 
allow more time for.SR to remove 



mean 

14 

was 



2- 



SO 



2- 



4 
Lakes 



Lake 



yielding a higher S0 4 retention, 
with a short residence time, like 
Anna, had a low S0 4 ^ retention (<10% of 
input) since the S0 4 is rapidly flushed 
out of the lake. The values of k su if i° 
Lake Anna (12-14 m/yr) were almost two 
orders of magnitude greater than the 
values reported by Baker et al . (1986). 
The implications of this drastic 
difference could be attributed to 
different S0 4 removal mechanisms. The 
lakes sampled by Baker et al . were all 
affected by acid precipitation, not AMD, 
and diffusion was said to be the major 
mechanism for S0 4 transport into the 
sediments. In Lake Anna, diffusion could 
account for no more than 5% of the 
observed S0 4 removal from the lake 
(Herlihy et al . 1987). A mechanism with a 



more rapid SO 



4 



account for the higher k, 



transport rate could 



observed in 



JUlf l 

Lake Anna. It has been hypothesized that 
S0 4 is transported to the sediment via 
adsorption onto settling solid particles 
or iron floe (Mills et al . in press). 



267 



Therefore the higher iron concentrations 
in Lake Anna could account for the higher 

k sulf- Tne value of k sulf in Lake Anna 
needed to make the model results match the 
observed results supports the hypothesis 
that some mechanism operating much more 
rapidly than diffusion is working to 
transport S0 4 to the sediments. 



Another 
model would 
coefficient 
concentratio 
the major S 
it is impo 
process in 
S0 4 remova 
of Contrary 
the iron pre 



possible improvement to the 
be to adjust the S0 4 removal 
to account for the iron 
n. If iron flocculation is 
4 transport mechanism then 
rtant to account for this 
the model. It is likely that 
1 is most rapid near the mouth 

Creek where the majority of 
cipitates. 



CONCLUSIONS 



In order to accurately 
observed sulfate distribution i 
it was necessary to reduce 
vertical diffusion coefficien 
amount of Contrary Creek wat 
into the epilimnion. The stee 
gradient near the mouth of Cont 
effectively inhibits vertical m 
observed data showed a grea 
fluctuation at station C2 indie 
the actual pattern of creek 
vertical diffusion has a grea 
temporal variability. 



model the 
n Lake Anna 
both the 
t and the 
er flowing 
p chemical 
rary Creek 
ixing. The 
t deal of 
ating that 
inflow and 
t deal of 



The 
Lake Anna 
removal 
showed th 
could be 
equation 
Lake Ann 
magnitude 
where dif 
S0 4 2 tr 
some oth 
important 
Anna . 



predict 

was ve 

coef f ic 

at the 

calcul 

3. Th 

a was 

highe 

fusion 

ansport 

er tran 

role i 



2- 



ed S0 4 concentration in 
ry sensitive to the S0 4 
ient^ The model results 
S0 4 removal coefficient 
ated successfully using 
e removal coefficient in 

more than an order of 
r than observed in lakes 
is the major mechanism of 

to the sediments. Thus 
sport mechanism plays an 
n removing S0 4 ~ from Lake 



Herlihy, A.T. 1987. Sulfur dynamics in an 
impoundment receiving acid mine 
drainage. Ph.D. Dissertation. Dept. or 
Environ. Sci., University of Virginia. 
Charlottesville. 214 pp. 

Herlihy, A.T. and A.L. Mills. 1985. 
Sulfate reduction in freshwater 
sediments receiving acid mine drainage. 
Appl. Environ. Microbiol. 49:179-186. 

Herlihy, A.T. and A.L. Mills. 1986. The 
pH regime of sediments underlying 
acidified waters. Biogeochemistry 
2: 95-99. 

Herlihy, A.T., A.L. Mills, G.M. 
Hornberger, and A.E. Bruckner. 1987. 
The importance of sediment sulfate 
reduction to the sulfate budget of an 
impoundment receiving acid mine 
drainage. Water Resour. Res. 23:287- 
292. 

Lung, W.S. 1987. A lake acidification 
model: a practical tool. J. Environ. 
Eng. In Press. 

Mills, A.L. 1985. Acid mine waste 
drainage: microbial impact on the 
recovery of soil and water ecosystems, 
pp. 35-81. Iq, D. Klein and R.L. Tate 
(eds.). Soil reclamation processes. 
Marcel Dekker Inc., New York. 

Mills, A.L.. and A.T. Herlihy. 1985. 
Microbial ecology and acidic pollution 
of impoundments, pp. 169-189. In» D. 
Gunnison (ed.), Microbial processes in 
reservoirs. Dr W. Junk Publ. , 
Dordrecht. 

Mills, A.L., P.E. Bell, and A.T. Herlihy. 
In press. Microbes, sediments and 
acidified waters: The importance of 
biological buffering. Id S.S. Rao (ed.) 
Microbial Interactions in Acid Stressed 
Aquatic Ecosystems. CRC Press, Inc., 
Boca Raton, FL. 



LITERATURE CITED 



Baker, L. A. . 
Pollman . 
alkalinity 
retention 



P.L. Brezonik, 
1986. Model of 

generation: 
component. Water 



and CD. 

internal 

sulfate 

Air Soil 



Pollut. 31:89-94 



Bruckner, A. E. 1986. Groundwater lake 
interactions in fractured rock terrane. 
M.S. thesis. Dept. of Environ. Sci., 
Univ. of Va., Charlottesville. 88 pp. 

DiToro, D.M. , J.J. Fitzpatrick, and R.V. 
Thomann. 1983. Water quality analysis 
simulation program (WASP) and model 
verification program (MVP) documenta- 
tion. Report submitted by Hydroscience, 
Inc. to the U.S. EPA environmental 
research laboratory, Duluth, MN. 



268 



PENNSYLVANIA'S APPROACH TO CUMULATIVE HYDROLOGIC IMPACT ASSESSMENT 
OF COAL MINING ACTIVITIES 1 



Lynn E. Langer 2 



Abstract .--Pennsylvania' s regulations require 
that an assessment be made of probable cumulative 
hydrologic impacts of all anticipated coal mining in 
the general area of a proposed mining operation and 
that the proposed operation be designed to prevent 
damage to the hydrologic balance. The cumulative 
hydrologic impact assessment (CHIA) process, there- 
fore, consists of two steps: (1) definition of 
potential damage to the hydrologic balance, and (2) 
prevention of such damage. Damage to the hydrologic 
balance is referenced to quality, quantity, and pres- 
ent uses of surface and ground water systems. The 
Pennsylvania Department of Environmental Resources ' 
(DER's) approach is to describe existing hydrologic 
conditions and potential adverse effects from mining 
in two-phase reports on watersheds of 20 to 50 mi 2 . 
The Phase I CHIA Report is a brief description of the 
watershed and a synopsis of surface and ground water 
uses and present and potential impacts from mining. 
The Phase II CHIA Report is a more detailed discussion 
of geology, hydrology, and mining history in the 
watershed. DER has found that insufficient data exist 
to adequately define existing hydrologic conditions. 
The phase approach allows available core information 
to be disseminated on a large number of watersheds 
while hydrologic data collection for more detailed 
assessment progresses. Prevention of damage to the 
hydrologic balance is accomplished primarily through 
evaluation of individual proposed mine sites. Because 
Pennsylvania's cumulative impact assessment program is 
based on existing permit review mechanisms, a permit 
reviewer can perform a mine site assessment even when 
a Phase I or II CHIA for the appropriate watershed has 
not yet been written. The major challenge which Penn- 
sylvania faces in implementing its CHIA program is in 
assuring the equitableness of effluent restrictions 
within individual watersheds. 



INTRODUCTION 

In order to obtain primacy in regu- 
lating its coal mining industry, Pennsyl- 



1 Paper presented at the 1988 Mine Drain- 
age and Surface Mine Reclamation Conference 
sponsored by the American Society for Sur- 
face Mining and Reclamation and the U.S. 
Department of the Interior (Bureau of Mines 
and Office of Surface Mining Reclamation 
and Enforcement), April 17-22, 1988, Pitts- 
burgh, PA. 

2 Lynn E. Langer is a Mining Specialist, 
Pennsylvania Department of Environmental 
Resources, Harrisburg, PA. 



vania was required to make numerous changes 
in its existing regulations. One change 
was the addition of a requirement that the 
regulatory agency (in this case, the Penn- 
sylvania Department of Environmental Re- 
sources ) evaluate each mining permit ap- 
plication in light of all existing and 
proposed mining operations in the general 
area and make an assessment of the probable 
cumulative hydrologic impacts. Proposed 
mining activities must be designed to pre- 
vent damage to the hydrologic balance with- 
in and outside the proposed permit area (25 
Pa. Code Section 86.37(a)(4), which is 
based on 30 CFR Section 773.15(c)(5)). 
This Pennsylvania regulation went into 
effect in 1982. 



269 



The cumulative hydrologic impact as- 
sessment, or CHIA, process consists of two 
steps: (1) definition of damage to the 
hydrologic balance, and (2) prevention of 
such damage. Under Pennsylvania regula- 
tions, damage to the hydrologic balance is 
referenced to quality and quantity of sur- 
face and ground water systems and to the 
present uses of the surface and ground 
water. Because these conditions vary from 
area to area, the definition of damage to 
the hydrologic balance does also. Accord- 
ingly, the first step of the CHIA process 
is to delineate the cumulative impact area 
and describe the current physical and hy- 
drologic conditions of that area. Existing 
and proposed surface and ground water uses 
are defined, and the susceptibility of 
those uses to potential adverse effects 
from mining activities is evaluated. Pre- 
vention of damage to the hydrologic balance 
(the second step of the CHIA process) in- 
volves a prediction of potential impacts 
from mining activities and efforts needed 
to prevent adverse impacts. 

In developing a CHIA program, DER has 
expanded on its own existing mechanisms for 
accomplishing both steps of the cumulative 
hydrologic impact assessment process. The 
CHIA has become a set of written documents 
wherein damage to the hydrologic balance is 
defined primarily in two-phase reports on 
watershed areas and efforts to prevent hy- 
drologic damage are addressed primarily in 
written assessments for individual mine 
sites . 



BACKGROUND 

Pennsylvania's environmental regula- 
tions require cumulative hydrologic impact 
analyses for six types of mining activi- 
ties: surface and underground mining of 
bituminous coal, surface and underground 
mining of anthracite coal, disposal of coal 
refuse, and operation of coal preparation 
plants. The Pennsylvania Department of 
Environmental Resources (DER) is the regu- 
latory agency responsible for reviewing 
permit applications and, consequently, for 
making impact analyses. Because DER re- 
ceives more permit applications for bitu- 
minous surface mines than for other types 
of coal activities, the agency has been 
orienting its initial efforts in develop- 
ment and implementation of the CHIA program 
toward bituminous surface mining, which 
will be the main thrust of this paper. 

Although Pennsylvania's formal program 
for cumulative hydrologic impact analysis 
is in the first stages of implementation, 
elements of CHIA have existed in the permit 
review program for a number of years. 
These elements include: (1) statewide 
designation of "Special Protection" water- 
sheds, where waste discharges of any nature 
are restricted; (2) mining management 
plans for sensitive watersheds; and (3) re- 
view of each proposed mine site for hydro- 
logic interactions with adjacent mines. 



Pennsylvania designates streams having 
excellent quality water and high public re- 
source value as Special Protection streams, 
specifically "Exceptional Value Waters" and 
"High Quality Waters". Regulations prohib- 
it degradation of ambient stream water 
quality in Exceptional Value Waters, which 
virtually eliminates further mining in 
those watersheds. High Quality Waters may 
be degraded from ambient quality if the 
discharge is a result of a project having 
public value and if downstream water uses 
will be protected. Mining operations pro- 
posed for High Quality watersheds have been 
required to meet more stringent permit re- 
quirements and have often been restricted 
in extent and nature. For several years, 
DER has required discharges from mining 
operations on High Quality watersheds to 
meet volume restrictions as well as efflu- 
ent quality standards in order to provide 
additional protection to the instream water 
quality. 

In the early 1980 's, DER developed 
mining management plans for watersheds 
where adverse cumulative impacts from min- 
ing were identified. Most of these water- 
sheds were public water supply sources. 
Typically, the mining management plans re- 
quired stricter erosion and sedimentation 
control measures, a limit on disturbed 
acreage on each mine site, mandatory over- 
burden analysis, and prohibition from min- 
ing coal seams which were known to produce 
acid mine drainage in the watershed. Many 
of these management plans are still being 
implemented. 

CHIA elements have long been part of 
the individual mining permit review pro- 
cess. Permit reviewers consider the inter- 
action of the proposed mining activities 
with adjacent active and abandoned mine 
sites. Typical analyses done by reviewers 
include determining probable impacts on the 
quality and quantity of discharges from 
underlying underground mines, impacts on 
surface flow and ground water recharge 
on adjacent surface mines, and cumulative 
impacts of several sites on ground water 
recharge to private water supplies or to 
wetlands. In recent years, DER has used 
effluent volume restrictions to address 
cumulative impacts on stream quality in 
some non-High Quality watersheds . 

The formalized CHIA process incor- 
porates and expands on the existing permit 
review elements. A standard format for 
impact analysis was developed to improve 
consistency of permit review among the 
various District Offices and help ensure 
that no aspect of cumulative impact assess- 
ment is overlooked. The resulting written 
format for analysis of impact assessments 
will also facilitate oversight by the fed- 
eral Office of Surface Mining Reclamation 
and Enforcement. Formalization of the CHIA 
process has ensured a systematic examina- 
tion of hydrologic impacts by watershed and 
has revealed a need for an expanded hydro- 
logic data base. 



270 



CHI A REPORT ELEMENTS 

Phase I CHIA Report 

Phase I provides a brief introduction 
to a cumulative impact area, highlighting 
areas of concern. For this part of the 
CHIA process, DER has chosen to use water- 
sheds of 20 to 50 mi 2 as cumulative impact 
areas. Each Phase I report is a short 
description of a watershed and a synopsis 
of surface and ground water uses, coal 
seams being mined, and previous and poten- 
tial hydrologic impacts of mining. For 
permit reviewers, it can increase famil- 
iarity with the area and encourage a water- 
shed-scale perspective. It alerts coal 
operators and consultants to potential 
impacts they might be expected to address 
in mine permit applications. 

Phase I includes the following sub- 
topics: (1) Hydrologic Unit, (2) Drainage 
Characteristics, (3) Physiography/Topogra- 
phy, and (4) Special Considerations. The 
Hydrologic Unit section defines the limits 
and location of the watershed and delin- 
eates it on a map. Drainage Character- 
istics include basin and subbasin drainage 
areas, maximum relief, stream lengths, and 
stream classifications. Physiography/Topo- 
graphy is a brief, general discussion of 
the physical characteristics of the water- 
shed. These three sections set the stage 
for the core of the Phase I report, Special 
Considerations . 

Special Considerations include a wide 
variety of hydrologic concerns. The pres- 
ence of a public water supply in the water- 
shed is often a major concern. Phase I 
describes the supply's location, source, 
and service area and any previous mining- 
related problems. Private water supplies 
(usually wells or springs) are common in 
the rural areas where bituminous surface 
mining takes place. Because their specific 
locations and sources are identified in 
permit applications, little emphasis is 
placed on reiterating this information in 
Phase I . Phase I does , however , note any 
previous adverse impacts on private water 
supplies in the area, such as diminution or 
degradation. 

As part of the Special Considerations, 
Phase I identifies water-based recreational 
uses which could be adversely affected by 
mining in the watershed. The most common 
instance is trout fishing in streams 
stocked by the Pennsylvania Fish Commis- 
sion. The use of wetlands, streams, and 
lakes by threatened or endangered species 
is also noted in this section. 

A brief summary of mining history and 
resultant hydrologic impacts in the cumula- 
tive impact area completes the Phase I re- 
port. This section identifies the coal 
seams which have been mined in the water- 
shed and whether they have produced acid 
mine drainage. A summary of the presence 
of toxic strata, calcareous strata, channel 
sandstones, and other strata of interest is 



made from overburden analyses performed in 
the area. Existing mining management plans 
are outlined. The report identifies previ- 
ous impacts which mining has had on water- 
shed resources and potential future hydro- 
logic impacts from mining. 

The Phase I CHIA is based on informa- 
tion from a number of sources. Reports by 
State and Federal agencies on regional coal 
hydrology and water resources are helpful 
for a broad overview, but they generally 
deal with much larger basin areas than the 
CHIA reports. For information which is 
more specific to the cumulative impact 
area, interviews with mining permit review 
and inspection staff personnel and review 
of DER data bases, biological stream sur- 
veys, and water guality reports for aban- 
doned mine reclamation projects are useful. 

Phase II CHIA Report 

In addition to describing vegetation, 
land uses, and soil types in the cumulative 
impact area, Phase II discusses watershed 
geology, hydrology, and mining history in 
much more detail than Phase I. 

Pennsylvania's initial attempt to 
write a Phase II CHIA revealed that exist- 
ing water quality and quantity data at key 
stream points are insufficient to adequate- 
ly define watershed hydrology. Most exist- 
ing data fall into two categories: (1) 
data from stream points with drainage areas 
larger than CHIA units (such as Pennsylva- 
nia Water Quality Network stations or USGS 
gauge stations ) ; or ( 2 ) data from coal op- 
erators ' water monitoring reports, which 
include stream points located immediately 
upstream and downstream of mine sites, but 
not necessarily at the receiving stream's 
mouth. In the 20 to 50 mi 2 watersheds 
which make up Pennsylvania ' s CHIA units , 
water quality and quantity data from major 
tributary confluences and critical points 
along the main stem (often including the 
main stem mouth) are minimal or absent. 

DER is currently using two approaches 
to develop the needed data on watershed hy- 
drology -- data collection by DER personnel 
and data collection by the United States 
Geological Survey (USGS). Bureau of Mining 
and Reclamation staff members are collect- 
ing quarterly water quality and quantity 
data from key points (main stem mouth, 
major tributary confluences, and other 
points) in two watersheds. Limited staff 
time prevents the use of this approach for 
additional watersheds. Therefore, DER is 
planning to initiate a joint project with 
the USGS to collect hydrologic data from 
additional watersheds. The project will be 
patterned after data collection by USGS on 
three prototype watershed projects in Penn- 
sylvania which are being funded by the Of- 
fice of Surface Mining Reclamation and En- 
forcement. Data collection on these pro- 
jects includes (1) continuous monitoring of 
the main stem mouth for stream discharge, 
pH, specific conductance, and temperature; 
(2) monthly water quality sampling of the 



271 



main stem mouth for additional parameters; 
and (3) four sets of water quality samples 
and flow measurements from key watershed 
points taken during baseflow and high flow 
periods . 

DER is also taking steps to increase 
the amount and availability of site- 
specific hydrologic data. Since November 
1986 the Department has required surface 
mine operators to measure, rather than 
estimate, the flows of all stream and 
discharge points they monitor, thus in- 
creasing the amount of reliable water quan- 
tity data available. In addition, the 
Department has completed compilation of 
existing quality and quantity data on acid 
mine drainage discharges from surface mines 
and is entering the results into a computer 
data base. This data base, categorized by 
drainage basin, will provide information on 
the severity and hydrologic impact of each 
mine drainage discharge, identification of 
the coal seams mined on the associated 
site, and a mechanism for prioritization of 
watersheds for further evaluation. 

The hydrology, mining history, and ge- 
ology sections are expected to be the most 
useful for permit reviewers and others in- 
terested in the CHIA. These sections will 
serve to document current "baseline" water- 
shed conditions for comparison with future 
conditions, detail the impacts previous 
mining has had on the hydrologic balance, 
and relate those conditions and impacts to 
the geology of the area. The vegetation, 
land use, and soils sections will increase 
in importance if DER finds computer model- 
ing of impacts to be valuable and practi- 
cal. 



Potential cumulative hydrologic im- 
pacts are as varied as the hydrologic con- 
ditions on potential mine sites. One fac- 
tor common to all sites, however, is the 
potential for impact on the quality of the 
receiving stream immediately downstream of 
the site. 

Under Pennsylvania's environmental 
regulations, every stream in the state is 
protected for certain minimum uses (potable 
water supply, fishing, esthetics, etc.) and 
also for specific additional uses, accord- 
ing to its classification. Therefore, even 
though a stream is not currently being used 
for a specific purpose (like those identi- 
fied in the Phase I CHIA Report), instream 
water quality may not be degraded beyond 
certain criteria. As previously discussed, 
the Department has been using discharge 
volume restrictions as a tool for prevent- 
ing violations of the instream criteria in 
High Quality watersheds and has been ex- 
panding their use to non-High Quality 
watersheds. 

At present, the volume restriction 
calculations are based on the acreage 
disturbed by the mine site and the total 
drainage area upstream of the discharge 
point. The incorporation of factors such 
as slope, soils, and vegetation through 
computer modeling is being explored by DER 
in a joint project with the U.S. Geological 
Survey. Implementation of computer model- 
ing in the CHIA program will be dependent 
on the amount of data collection required 
for accurate calibration of the model and 
on the transferability of the calibrated 
model from one watershed to another. 



If current mining management practices 
are found to be inadequate to prevent cumu- 
lative damage to the hydrologic balance, 
the Phase II report will also develop rec- 
ommendations for future management strat- 
egies. 

Mine Site Assessment 

The Mine Site Assessment is a written 
appraisal of the probable impacts of a spe- 
cific site on the hydrologic balance, in 
conjunction with existing and anticipated 
mine sites in the general area. The Mine 
Site Assessment describes the proposed min- 
ing activities and how they have been 
planned to prevent damage to the hydrologic 
balance . 

In evaluating potential damage to 
existing uses of surface and ground water, 
the permit reviewer relies on hydrologic 
data and resource information from several 
sources, primarily the permit application 
and the Phase I and II CHIA Reports. Be- 
cause of the information available in the 
permit application, the inspector's field 
review report, and reports submitted by 
other agencies, a permit reviewer can per- 
form a Mine Site Assessment even if the 
Phase I and II CHIA Reports for that water- 
shed have not yet been written. 



CHALLENGES IN IMPLEMENTATION 

The major challenge which Pennsylvania 
faces in implementing its CHIA program is 
in the area of assigning effluent quality 
and quantity limits. DER strives to be 
equitable in the restrictions it places on 
operations within a single watershed. 
Three approaches exist to accomplish this: 
(1) base effluent limits on the maximum 
acreage predicted to be disturbed in the 
watershed at one time; (2) impose new, 
stricter limits on existing operations when 
additional discharges are approved; or (3) 
establish an upper limit of disturbed 
acreage for each watershed, and require new 
operations to wait until older ones are 
completed. 

For the first approach, it is virtual- 
ly impossible to predict the maximum acre- 
age which will be disturbed at one time in 
any given watershed. An estimate of re- 
maining mineable coal reserves could be 
substituted, representing the absolute max- 
imum possible, but this tactic would re- 
quire significant amounts of information 
not readily available to the Department. 
In addition, it could result in measures 
overly restrictive for protection of the 
water resources. 



272 



The second approach, imposing more re- 
strictive effluent limits on existing oper- 
ations, may require substantial physical 
modifications on those sites which might 
not be technologically or economically fea- 
sible. 

The third approach, establishing a 
ceiling on the amount of disturbed acreage, 
involves considerable administrative effort 
and may impose economic hardships and plan- 
ning difficulties on individual mine opera- 
tors. 



CONCLUSIONS 

The Pennsylvania Department of Envi- 
ronmental Resources has attempted to devel- 
op a practical approach to cumulative hy- 
drologic impact assessment. Existing in- 
formation on surface and ground water uses 
and on past and potential hydrologic im- 
pacts from mining is made readily available 
to permit reviewers through Phase I CHIA 
Reports. Additional key water quality and 
quantity data are being collected and will 
be made available, along with detailed ge- 
ology and mining history discussions, in 
the Phase II CHIA Reports. The two-phase 
report allows basic, core information to be 
disseminated on a large number of water- 
sheds, facilitating permit review, while 
more detailed data collection progresses. 
Reviewers of mine permit applications use 
all available information to make assess- 
ments of cumulative hydrologic impacts and 
to ensure that individual mining operations 
are designed to prevent damage to the hy- 
drologic balance. 



273 



BIOLOGICAL METAL REMOVAL FROM MINE DRAINAGE 



Katsutaka Nakamura" 



Abstrac 
metals requi 
surface wate 
on known bio 
cost-ef f ecti 
of metal in 
with a conti 
reactor, wit 
ous flow sys 
principal su 



t. — 
res 

rs . 
logi 
ve p 
mine 
nuou 
h a 
tern 
lfat 



Mine drain 
proper tre 

The aim o 
cal method 
ilot plant 

drainage . 
s-flow flu 
working vo 
using Desu 



age co 
atment 
f this 
s and 
-scale 
Good 
idized 
lume o 
If ovib 



e-reducing bacte 



ntaminated with heavy 
to avoid pollution of 
work was to improve 

to develop an economic, 
process for removal 
results were obtained 
bed-type anaerobic 

f 180L, in a continu- 

rio vulgaris as the 

ria. 



INTRODUCTION 

There are a number of inactive and 
abandoned underground mines in Japan from 
which acid drainage with heavy metals flows 
out continuously. To alleviate the acidic 
condition and metal content in the drainage, 
the following chemical treatment has been 
carried out: neutralization by slaked lime 
and powdered limestone. This chemical 
treatment, however, is not economically 
attractive because of its high cost and the 
occurrence of a large quantity of waste 
sludge. Therefore, alternative methods 
are needed. 

Since dissolved metals can be removed 
as sulfide precipitates by anaerobic sul- 
f ate-reducing bacteria, various attempts 
have been made to treat acid mine drainage 
by this biological methods (Wakao, et al . 
1979, Maree and Strydom 1985, 1987). 

In this work basic experiments were 
carried out using Desulf ovibrio vulgaris in 
both laboratory and pilot-scale tests in a 
continuous flow system. In developing the 
process, the following aspects were inves- 
tigated : 
(1) Optimum conditions of a process to 
maintain the biological and chemical 
reactions: amount of organic carbon, 
pH level, inflow of raw water, and 



1 Katsutaka Nakamura is Deputy Director, 
Mine Pollution-Control Technical Develop- 
ment Division, Technical Development 
Department, Metal Mining Agency of Japan. 



temperature for reactions 

( 2 ) Type of reactor required for effective 
separation of sulfide precipitate 

(3) Continuous process to remove metal 
from raw water. 



LABORATORY-SCALE EXPERIMENT 

A strain of Desulf ovibrio vulgaris was 
placed in culture medium shown in table 1 
and incubated at 30 C. After incubation 
for six days, the bacteria grew and the 
medium turned black in color. The degree 
of reduction of sulfate to sulfide is shown 
in figure 1. This enriched culture was 
used as an inoculum of the culture of sul- 
f ate-reducing bacteria used in this experi- 
ment . 

The experiment was performed in a con- 
tinuous-flow fluidized bed-type anaerobic 
reactor with a working volume of 10L. The 
water used as feed was synthetic, composed 
of the same elements as shown in table 1. 
The water was fed continuously to the re- 
actor after sufficient growth of the bac- 
teria. Hydrogen sulfide was bubbled into 
the solution for mixing and coagulant, 
polyacrylamide, was added in the reactor to 
step up precipitation during the run. 

Both the Fe2+ and Zn 2 + ion contents in 
the feed water were set at 50mg/L respec- 
tively. Quantity of these metal ions and 
the sulfate concentration value in the ef- 
fluent was used as the indicator of the re- 
actions . 

The experiment was performed changing 



274 



Table 1 . --Chemical characteristics of the 
culture medium used in the incubation 
of bacteria and laboratory-scale ex- 
periment . 



Component 



K2HP04 

NH4C1 

Na2S04 

MgS04-7H20 

Na-lactate 

Yeast extract 

FeS04 • (NH4 ) 2S04 • 6H20 

Cu2+CCuS04-5H20] 

Pb2+CPb(N03)2] 

Zn2+CZnS04-7H20] 

Cd2+CCdS04-nH20] 

As3+CNaAs02D 
Tap water 
pH 



Value 



0.5g 



1 


g 


1 


g 


2 


g 


3. 


5g 


1 


g 


0. 


5g 


3 


g 


10 


mg 


50 


mg 


1 


mg 


1 


mg 


1,000 


mL 


6.0-6.5 





Table 2. --Result of the laboratory-scale 
experiment . 



Inflow 
(L/day) 


Retention 
time 
(hours) 


Temp. 
(°C) 


Effluent 


pH 


Zn2+ 
(mg/L) 


10 


24 


30 


6.21-8.01 


0.09-0.82 


20 


12 


30 


6.19-6.53 


0.14-0.30 


40 


6 


30 


6.14-6.78 


0.08-1.47 


80 


3 


30 


6.23-7.00 


0.10-0.68 


40 


6 


20 


6.20-7.00 


0.03-0.31 


40 


6 


15 


6.50-6.71 


0.13-0.18 


80 


3 


15 


6.40-6.60 


0.16-0.26 



(3) Concentration of the bacteria typical- 
ly ranged from about 4-5X10^ cells/mL 
throughout the experiment. 



PILOT PLANT-SCALE TEST 



(mg/L) 

c 2000i 

o 



1500 



1000- 



500 



Occurrence of sulfides 



-1 1 1 r- 



Time (days) 

Figure 1. --Degree of decrease of sulfate 
by sulf ate-reducing bacteria, 
Desulf ovibrio vulgaris . 

inflow stepwise at 10, 20, 40, and 80L/day, 
with the pH level between 5.0 and 6.5 and 
temperature in the reactor at 15 C and 30 C. 



s the organic 
rst at 3.5g/L and 
tract was started 
finally, 
the feed water 
consistently 
inuous operation: 
ent was kept be- 
in Japan ( Fe < 10 
, Pb <lmg/L, Cd< 



Sodium lactate used a 
carbon source was added fi 
reduced to lg/L. Yeast ex 
at lg/L, then cut to zero 

Most of the metals in 
were separated as sulfides 
during the periods of cont 
metal content in the efflu 
low the emission standard 
mg/L, Zn<5mg/L, Cu < 3mg/L 
O.lmg/L, As<0.5mg/L). 

Concentration of the bacteria in the 
reactor was measured by the Thoma-Zeiss 
counting chamber once a week. 

The results of the experiment was 
shown in table 2 and was summarized as 
follows . 

(1) Sodium lactate can be cut down to lg/L 
at 20°C and pH5, and to 2g/L at 10°C 
and pH6 

(2) Retention time was 3 hours ( 80L/day in- 
flow) at 15 C, pH6, sodium lactate 

3 . 5g/L and yeast extract lg/L, and 6 
hours ( 40L/day inflow) at pH5 



Materials and Methods 

A pilot plant-scale test was performed 
at Yanahara Mine in the Okayama Prefecture, 
based on the results of the laboratory ex- 
periment . 

The test was in a continuous flow pro- 
cess using a fluidized bed-type anaerobic 
reactor with a working volume 180L, which 
was scaled-up version of the laboratory 
experiment system. 

A schematic diagram of the plant and 
flow diagram of the process used are illus- 
trated in figure 2 and figure 3, respec- 
tively . 

The culture of the bacteria used for 
the test was the same as the previous labo- 
ratory experiment. 

Mine drainage of Yanahara Mine, the 
quality of which is shown in table 3, was 
supplemented with sodium lactate, yeast 
extract, and inorganic materials. This 
mixed water was fed to the anaerobic re- 
actor . 

As the raw drainage was too acidic (pH 
2.50-2.65) to maintain the activity of the 
bacteria and to clear the emission standard, 
the drainage was controlled by adding pow- 
dered limestone, the average grain size of 
which was 1.1 micrometer, to maintain pH at 
about 5 before the mixing. The condition 
of the influent is shown in table 4. 

Table 3. --Chemical characteristics of mine 
drainage from the Yanahara Mine used 
for the pilot-plant scale test. 



Component 



Value 



temperature 
pH 
Fe2 + 

Total-Fe 
Zn2 + 
Cu2 + 
Cd2 + 
Mn2 + 
Total-As 
Acidity as CaC03 



20°C 

2.5-2.65 

171-303 mg/L 

693-815 mg/L 

41.5-52.1mg/L 

9.4-16.8mg/L 

0.18mg/L 

7.4 mg/L 

0.21mg/L 

1, 991-2, 442mg/L 



275 



Table 4 . --Condition of the feed water used 
throughout the pilot plant-scale test. 



Temperature 
(°C 


Inflow 
(L/day) 


Yeast 
extract 

(g/L) 


Na- 
lactate 

(g/L) 


P H 


Period 
(days) 


24.7-30.2 


180 


0.5 


3.5 


6 


16 


25.8-28.8 


360 


0.25 


3.5 


6 


2 


22.2-27.7 


360 





3.5 


5 


12 


19.0-25.6 


720 





3.5 


5 


13 


18.1-22.8 


720 





3 


5 


8 


14.8-19.8 


720 





2.5 


5 


10 


10.0-16.2 


720 





2 


5 


20 


10.8-14.9 


720 





1.5 


5 


18 


9.0-14.2 


720 





1 


5 


14 



Raw drainage 



The test was made increasing feed 
volume at three steps, 180, 360 and 720 
L/day and was performed for 113 days conti- 
nuously. 

Sodium lactate and yeast extract were 
reduced from 3.5g/L to l.Og/L and 0.5g/L 
to zero, respectively. Temperature in the 
reactor ranged between 9.0 C and 30.2°C, 
depending on the weather, because the test 
plant was not housed in order to test the 
procedure in practical use too. 

Excess hydrogen sulfide except for the 
volume, which was bubbled into the reactor, 
was vented. 

The COD in the effluent was measured 
by the dichromate titration. 



Coagulant stock tank 




Level controller 



N 2 gas . slower 
cylinderl 



Figure 2. --Diagram of drainage treatment system used in the pilot plant-scale test. 



Yeast extract 



Na-lactate 



Inorganic materials 



NH CI 
4 



Na o S0„ 
2 4 



Nutrient solution 



Raw drainage 


CaCO 




Neutralizing 








< 


' 






Drainage 

2+ 
Fe 13-95 mg/L 

2+ 
Zn 36-59 mg/L 

pH 4.7-5.5mg/L 



Influent 


water 






K.HPO. 

2 4 




0.5 


g/L 


NH CI 
4 




1 


g/L 


Na.SO„ 
2 4 




1 


g/L 


MgS0 4 ■ 


7H 


2 


g/L 


Na-lactate 


3.5 


1 g/L 


Yeast extract 


0.5 


g/L 


Fe 2+ 




10-76 


mg/L 


Zn 2 + 




29-47 


mg/L 


PH 




4.7-5.8 





Figure 3. --Flow diagram of feed water in the pilot plant-scale test. 



276 



Fe 



2 + 



(mg/L) 
80 

70 

60 -\ 

50 

40- 

30 - 

20 - 

10 





(mV) 
100 



Eh 



-100' 



-200 



Yeast 
extract |- 
(g/D 

Na-lactate , 
(g/L) r 

Retention. 



time 
(hours) 



influent water 
ill I 



I V- ^ 



'' » i * i \ 

i ■ ' ■ > 

« ! ' M 

,\ a i i '< 

lUiy 1 



A 



M*N 



\ 



\ 






effluent water 
1 ' 



^s ^* — 



V 



VJ 



AN 




0.5 



0.25 
I I 



3.5 



2.5 



1.5 



1.0 



24 



12 



— l — 
20 



40 60 

Time (days) 



-i — 
80 



100 



Figure 4. --Result of pilot plant-scale test. 



120 



Results 

Most of the metals in the mine drain- 
age were separated into precipitate of sul- 
fides consistently as was the case in the 
laboratory experiment. The results of the 
test are shown in figure 4. The Fe2+ and 
Zn2+ ions in the feed water were reduced to 
Fe2+<1.0mg/L, Zn2+ < . 2mg/L, below the 
emission standards in Japan, except for the 
period of the start-up and the very last 
stage. The cause of the increasing Fe2+ in 
the effluent at the start-up is not known. 
The same phenomenon at the end of the ex- 
periment was clearly due to the decreasing 
of Na-lactate. The COD in the effluent, 
however, was 800-1 , OOOmg/L throughout the 
test, which is over the standard. 

The followings was concluded from the 
results of the pilot plant-scale test. 

(1) A cut in the amount of yeast extract 
was possible 

(2) Mine drainage with chemical quality of 
Fe2+ 40mg/L, Zn2+ 40mg/L and pH4.7 was 
treated successfully at 12 C, sodium 
lactate level at 1.5g/L, and retention 
time of 6 hours ( 720L/day inflow) 



(3) Concentration of the bacteria remained 
at from 4-5X109 cells/mL as same as 
was in the laboratory experiment. 



CONCLUSIONS 

A continuous flow test of a biological 
method for treatment of mine drainage at 
Yanahara Mine showed that most of the met- 
als in the mine drainage can be reduced in 
a pilot plant-scale process when suitable 
organic carbon for sulf ate-reducing bacte- 
ria is supplied, and inflow and temperature 
of the reactions are properly controlled. 

However, the following problems must 
be solved to complete a system for practi- 
cal use : 

(1) The COD value in the effluent must be 
decreased to within the emission 
standard 

(2) Cheaper nutrients must be selected in 
order to hold cost down. 



277 



ACKNOWLEDGEMENT 

This work was performed as one of the 
projects of research and development for 
mining-related pollution control by the 
Metal Mining Agency of Japan. The author 
wishes to thank all the members concerned 
with the project, particularly Dr. Imai, a 
professor of Okayama College of Commerce, 
for his presenting useful bacteria, 
Desulf ovibrio vulgaris . 



LITERATURE CITED 

Maree J. P. and Strydom Wilma F. (1985) 

Biological sulfate removal in an up- 
flow packed bed reactor. Wat. Res. 19, 
1101-1106. 

Maree J. P. and Strydom Wilma F. (1987) 

Biological sulfate removal from in- 
dustrial effluent in an upflow packed 
bed reactor. Wat. Res. 21, 141-146. 

Wakao N., Takahashi T. , Sakurai Y. and 

Shiota H. (1979) A treatment of acid 
mine water using sulf ate-reducing 
bacteria. J. Ferment. Technol . 57, 
No. 5, 445-452. 



278 



GROWTH RESPONSES AND IRON UPTAKE IN SPHAGNUM PLANTS 
AND THEIR RELATION TO ACID MINE DRAINAGE TREATMENT 1 



Anne Kearney Spratt and R. Kelman Wieder' 



Abstract.-- Increasing interest in the use of 
Sphagnum wetlands to treat acid mine drainage 
(AMD) has prompted study of the tolerance of 
Sphagnum species to the various constituents of 
AMD waters. In this study, S^ f allax and S . 
henryense plants were grown in the laboratory for 
33 days in synthetic bog water to which Feci, was 
added to achieve Fe 2+ concentrations ranging from 
to 10,000 mg/L. Upon exposure to Fe concentra- 
tions >^ 100 mg/L, both species exhibited a signi- 
ficant reduction in growth after 33 days relative 
to controls, but this reduction was proportion- 
ately less for S^ f allax than for S^ henryense . 
Final Fe concentration in the plant tissue was 
negatively correlated with growth and chlorophyll 
concentration in both species, and greatly 
exceeded the total cation exchange capacity for 
Sphagnum , suggesting nutrient cation deficiency as 
a mechanism of Fe toxicity. Based on data from a 
previous study of net primary productivity of 
Sphagnum species in Big Run Bog, estimates were 
made for Fe accumulation by the growing plants on 
an areal basis. Projected areal accumulation of 
Fe was greatest at high Fe treatment concentra- 
tions (1,000-10,000 mg/L) for both species, 
despite significant decreases in plant growth. 
These estimates reinforce the conclusion that Fe 
uptake by growing Sphagnum plants can play only a 
relatively minor role in Fe retention in wetland 
systems constructed for mine drainage treatment. 
Results of this study also indicate that the via- 
bility of Sphagnum wetlands constructed for AMD 
treatment may be dependent on the species com- 
position . 



■■■Paper presented at the 1988 Mine Drainage 
and Surface Mine Reclamation Conference 
sponsored by the American Society for 
Surface Mining and Reclamation and the u. 
S. Department of the Interior (Bureau of 
Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 
1988, Pittsburgh, PA. 

o 

'Anne Kearney Spratt is a research 

technician, Division of Pinelands 
Research, Rutgers University, New Lisbon, 
NJ, and R. Kelman Wieder is an associate 
professor, Department of Biology, 
Villanova University, villanova, PA. 



INTRODUCTION 

As a result of several studies 
suggesting that chemical and biological 
processes in freshwater wetlands may 
remove acidity, sulfate, and heavy metals 
from acid coal mine drainage (Wieder and 
Lang 1982, Kleinmann et al. 1983, Burris 
et al. 1984, Tarleton et al. 1984, Wieder 
and Lang, 1984, Gerber et al. 1985, Wieder 
et al. 1984a, 1985, McHerron 1986, Brodie 
et al. 1987), there has been an increasing 
interest in using constructed wetland 
systems as a low-cost, low-maintenance 
alternative to traditional chemical treat- 



279 



ment of acid mine drainage (AMD) . Before 


the wetland approach to AMD treatment 


becomes feasible, however, the effects of 


AMD on wetland vegetation, including 


Sphagnum, must be evaluated. The vegeta- 


tion in man-made wetland systems serves 


not only as a mechanism of metal retention 


by the process of plant uptake, but also 


as a mechanism of stabilization of the 


organic substrate within the wetland, 


minimizing the potential for erosion. 


Under natural conditions, Sphagnum 


growth and species distributions are 


affected by environmental factors such as 


pH, moisture, and nutrient availability 


(Clymo 1973, Vitt et al. 1975, Andrus et 


al. 1983, Titus et al. 1983, Wagner and 


Titus 1984, Andrus 1986). Sphagnum 


species can accumulate Fe and other metals 


with no apparent inhibition of growth when 


subjected to chronic inputs of low concen- 


trations of metals via atmospheric deposi- 


tion (Pakarinen and Tolonen 1976, 


Pakarinen 1978, Aulio 1980, Lembrechts and 


Vanderborght 1985) . Accumulation under 


these conditions is species specific 


(Aulio 1982) and is not necessarily corre- 


lated with plant growth (Pakarinen and 


Rinne 1979) . 



treatment solution was added. Treatment 
solutions contained either 0, 1, 10, 100, 
1,000, or 10,000 mg/L Fe , thereby spanning 
the range of Fe concentrations typical of 
AMD (Braunstein et al. 1977). Ferrous 
chloride, rather than FeSO^, was used in 
preparing the treatment solutions. A 
previous study of the growth of S_. f allax 
and S. henryense in AMD waters demon- 
strated that growth was negatively corre- 
lated with SO^ , but not with Cl~, con- 
centrations in AMD (Kearney, 1986). The 
treatment solutions were prepared by 
adding Fed? to synthetic bog water con- 
taining, in mg/L: 0.88 Ca 2+ , 0.19 Mg 2+ , 
0.47 K , 0.23 Na + , 0.16_NH 4 , 0.07 NO3 , 
3.44 S0 4 2- , and 0.62 Cl~. The synthetic 
bog water approximates the ion concentra- 
tions found in Big Run Bog surface waters 
(Wieder 1985). Treatment solutions were 
added to each bowl so that the water table 
was at the tops of the PVC cylinders. 
Every 2-3 days, additional treatment solu- 
tion was added to maintain the water at 
the tops of the cylinders, and at weekly 
intervals the entire volume of water in 
each bowl was replaced. Plants were grown 
for 33 days beginning on December 23, 

1985, for S^ f allax and on January 8, 

1986, for S. henryense . 



In contrast to the considerable 
amount of information available on 
Sphagnum growth in relatively undisturbed 
environments, comparatively little is 
known about the response of Sphagnum to 
the low pH and high metal concentrations 
typical of AMD. One study has demon- 
strated that the discharge of chemically 
treated mine drainage (pH values from 6 to 
9) into what was once a naturally acidic 
Sphagnum wetland (surface water pH near 
4.0) resulted in both the death of the 
vegetation and erosion of the organic peat 
down to the underlying mineral soil 
(Wieder et al. 1984b) . 



ob je 

the 

occu 

with 

AMD, 

Fe f 

spec 

that 

coul 

land 

buti 

proc 



In 
ctiv 
grow 
rr in 

Fe 

and 
rom 
ies . 

upt 
d ma 

is 
on t 
esse 



light 
es of 
th re 
g Sph 



of thes 
this st 
sponses 
agnum sp 



conce 
to a 
solut 
Als 
ake o 
ke to 
compa 
hat o 
s cou 



ntrat ion 
ssess th 
ion by t 
o, the p 
f Fe by 

Fe rete 
red to t 
ther che 
Id make 



e find 
udy we 
of two 
ecies 
s typi 
e appa 
hese t 
otent i 
growin 
nt ion 
he pot 
mical 
to Fe 



ings , 
re to 

commo 
in sol 
cally 
rent u 
wo Sph 



the 

assess 
nly 

utions 
found in 
ptake of 
agnum 



al con 
g Spha 



tr ibut ion 
gnum 



within 
ential 
and bi 
retent 



a wet- 
contr i- 

ological 

ion . 



METHODS 



speci 

were 

Sphag 



Livi 
es , 
coll 

num- 



AMD. 

of ea 

lengt 

or 8 

cylin 

these 

diame 



In 
ch s 
h of 
into 
ders 

cyl 
ter 



ng pla 

Sphagn 

ected 

domina 

the la 

pecies 

5 cm 

2.5-C 

(Wied 

inders 

finger 



nts 
urn f_ 
at B 
ted 
bora 

wer 
and 
m-di 
er e 

wer 
bowl 



of two Sphagnum 
allax a 



lg Run 
wetland 
tory, i 
e cut t 
placed 
ameter , 
t al . 1 
e place 
to whi 



nd S 
Bog, 

not 
ndivi 
o an 
in gr 

3.5- 
984b) 
d in 
ch a 



henryense , 
WV, a 

affected by 
dual plants 
initial 
oups of 7 
cm-tall PVC 
Three of 
an 8-cm- 
particular 



Plant growth was assessed by removing 
individual plants from their cylinders and 
measuring their length with a metric ruler 
at 11-day intervals. Since growth of 
Sphagnum is indeterminate and occurs along 
the long axis of the plant, increase in 
length is an appropriate measure of growth 
over time (Clymo 1970) . 



chlor 

Using 

from 

tissu 

weigh 

the f 

aceto 

extra 

using 

chlor 

using 



were 
at 55 
deter 
was a 
for 1 
ash w 
to fi 
(Like 
trati 
deter 
photo 
ology 



Afte 
ophy 

app 
each 
e (c 
ed. 
resh 
ne ( 
ct a 

a d 
ophy 

equ 

The 
weig 
C 
mina 
shed 
hou 
as e 
nal 
ns a 
ons 
mine 
met r 
se 



r 33 d 

11 con 

roxima 

bowl , 

apitul 

Chlor 

plant 

Dolphi 

t 649 

ouble- 

11 con 

ations 

plants 
hed an 
for f r 
tions . 

in a 
r and 
xtract 
volume 
nd Bor 
in the 
d by a 
y . Fo 
e Kear 



ays , d 
centra 
tely h 

the u 
urn) wa 
ophyll 

tissu 
n 1978 
and 66 
beam s 
centra 

given 

remai 
d then 
esh ma 

The 
muffle 
at 800 
ed wit 

with 
mann 1 

extra 
tomic 
r more 
ney ( 1 



eterminat 
tion were 
alf of th 
ppermost 
s excised 

was extr 
e with 80 
) . Absor 
5 nm was 
pectropho 
tion was 

in Dolph 



ning 

dri 

ss/d 

dr ie 

fur 

C 

h 6 

dist 

970) 

ct s 

abso 

det 

986) 



in e 
ed f o 
ry ma 
d pla 
nace 
for 3 
M HC1 
illed 
Ir 
oluti 
rptio 
ailed 



ions of 

made . 
e plants 
1 cm of 

and 
acted from 
% aqueous 
bance of the 
measured 
tometer, and 
calculated 
in (1978) . 

ach bowl 
r 24 hours 
ss ratio 
nt material 



at 300" C 

hours. The 

and brought 

water 
on concen- 
ons were 
n spectro- 

method- 



For each species, effects of Fe con- 
centration in the treatment solution on 
growth (increase in length over the 33-day 
period) and chlorophyll concentration were 
assessed using Friedman's tests conducted 
on the data from day 33. Correlations 
between growth, final chlorophyll concen- 
tration, final tissue Fe concentration, 



280 



and Fe concentration in the treatment 
solutions were determined using Spearman's 
rank correlation tests. Differences 
between the species with regard to growth 
and final tissue Fe concentration were 
evaluated with Mann-Whitney tests. A 
significance level of p=0.05 was used in 
all tests. 



RESULTS 



both 

grow 

trea 

33 d 

was 

cont 

the 

sign 

solu 

trat 

plan 

1) . 

at a 

less 

(fig 

in 1 

f ica 



The 
simi 
th re 
tment 
ays g 
signi 
rol s 

10 mg 
if ica 
t ion . 
ions 

t gro 
In c 

11 Fe 
than 

. 2) . 
ength 
ntly 



two Sphagnum 

larities 

sponses 

solutio 
rowth in 
f icantly 
olution 
/L Fe so 
ntly fro 
Howeve 
increase 
wth prog 
ontrast , 

concent 

growth 
For ea 

achieve 
negative 



and 
to i 
ns . 

the 

gre 
(0 m 
luti 
m th 
r, a 
d fr 
ress 

for 
rati 
in t 
ch s 
d ov 
ly c 



speci 
diffe 

ncreas 
For S 
1 mg/ 

ater t 

g/D , 

on did 
at in 
s solu 
om 100 
ively 

S. he 
ons wa 
he con 
pecies 
er 33 
orrela 



es e 
renc 
ing 

L Fe 

han 

and 

not 
the 
tion 

to 
deer 
nrye 



xhibi 
es in 
Fe in 
llax, 



solu 
that 
growt 

diff 
contr 

Fe c 
10,00 
eased 
nse g 



s si 
trol 
, th 
days 
ted 



gnif i 
solu 

e inc 
was 

with 



ted 

the 

after 
tion 
in the 
h in 
er 
ol 

oncen- 
mg/L, 

(fig. 
rowth 
cantly 
tion 
rease 
signi- 
treat- 



ment 


solution 


Fe concentration (Spearman's 


rho 


values of 


-0.82 and -0.93 for S. 


fall 


ax and S. 


henryense, respectively). 


Also 


for both 


species, at solution Fe 


concentrations 


> 100 mg/L not only was 


grow 


th reduced 


relative to controls, but 


also 


no significant increase in growth was 


observed after 


the first 11 days. Regard- 


less 


of treatm 


ent solution Fe concentra- 


tion 


, the inhi 


bition of growth relative to 


cont 


rol plants 


was proportionately greater 


for 


S. henryense than for S. fallax 


(fig 


. 3) . 





Final chlorophyll concentration in S. 
fallax generally decreased with increasing 
treatment solution Fe concentration (fig. 
1) . For S_^ henryense , final chlorophyll 
concentration in treatment solution Fe 
concentrations _< 10 mg/L were not signifi- 
cantly different from those plants grown 
in mg Fe/L, but as Fe concentrations in 
treatment solutions increased to 100 mg/L 
and greater, chlorophyll concentration 
decreased (fig. 2). In each species, 
chlorophyll concentration in plant tissues 
after 33 days was positively correlated 
with increase in length (Spearman's rho 
values of 0.68 and 0.82 for S^ fallax and 
S. henryense, respectively). Moreover, at 



0.50 r 



E 
o 

£ 0.25 

S 

o 



0.00 




F* 

mg/L 
1 

10 



3- w.j 



100 



1000 



10000 



*,y 



y.* 



1 



0.0 0.8 1.6 2.4 

Final chlorophyll concentration 

(mg/g dry mass) 



Figure 1. — Growth and final chlorophyll concentration of Sphagnum fallax in Fe solutions. 
(For growth, each symbol represents the arithmetic mean of 72 values + 1 standard 
error. Values for mean length after 33 days with the same letter (a-e) do not differ 
significantly. For chlorophyll concentration, the mean and range of 3 measurements 
are shown. Values with the same letter (w-z) do not differ significantly. Iron 
concentrations in the treatment solutions are indicated to the left of the chlorophyll 
bars) . 



281 



0.50 



E 
u 

£ 0.25 

S 

o 





0.00 



S. henryens* 




mg/L 
o 



1 

10 



100 

1000 

10000 



3- 



w 



3- 



y>* 



10 



15 



20 



25 



30 



35 



0.0 



1.0 



2.0 3.0 



Days 



Final chlorophyll concentration 

(mg/g dry mass) 

Figure 2. — Growth and final chlorophyll concentration of Sphagnum henryense in Fe solu- 
tions. (For growth, each symbol represents the arithmetic mean of 63 values _+ 1 
standard error. For chlorophyll concentration, the mean and range of 3 measurements 
are shown. For explanation of letter symbols, see Figure 1 legend. iron concentra- 
tions in the treatment solutions are indicated to the left of the chlorophyll bars.) 



Fe concentrations J> 100 mg/L, final chlor- 
ophyll concentration was less than 1.2 
mg/g dry mass and no additional growth was 
achieved after the first 11 days. 



For both species, Fe concentration in 


the plant tissues after the 33-day period 


increased with increasing treatment solu- 


tion Fe concentration (fig. 4), yet there 


was no significant difference between the 


two species in final tissue Fe concentra- 


tion across all treatment solution Fe 


concentrations (p=0.34). For each 


species, growth achieved after 33 days was 


negatively correlated with final tissue Fe 


concentration (Spearman's rho values of 


-0.83 and -0.86 for S. fallax and S. 


henryense, respectively). 



DISCUSSION 

The observed reduction in Sphagnum 
growth with increasing Fe concentration in 
the treatment solutions may be related to 
a decreased ability of the plants to take 
up nutrient cations. Cation exchange on 
the cell walls of Sphagnum plants appears 
to be a major mechanism of cation uptake 
by the living cells of the plants (Clymo 
1967). The maximum reported values for 



120 r f-l 



90 



| 60 

£ 

JS 

% 

o» 30 



10 



100 1000 10000 



Solution concentration (mg Fe/L) 

Figure 3. — Mean growth after 33 days of 
Sphagnum fallax (open bars) and S. 
henryense (stippled bars) in Fe solu- 
tions, expressed as percent relative 
to control treatments (0 mg Fe/L). 



282 



240 



1S0 



I I 

g £• 120 



:s 



rU 



r-rh 



n^i 



i 



- 8 



[ill. 



1 10 100 1000 10000 

Solution concentration (mg Fe/U 

Figure 4. — Final concentrations of Fe in 
the tissues of Sphagnum f allax (open 
bars) and S_^ henryense (stippled 
bars) . The arithmetic mean and range 
of 3 measurements are shown. 



cation exchange capacity in Sphagnum 
species are close to 1 meq/g dry mass 
(Clymo 1967, Spearing 1972), and base 
saturation of Sphagnum (percent of total 
cation exchange capacity accounted for by 
the base cations Ca 2+ , Mg 2+ , K + , and Na + ) 
has been measured as 8.7% (Braekke 1981). 
Cation exchange is a dynamic process in 
which equilibria between adsorbed cation 
and soluble cation concentrations are 
rapidly established (Clymo 1963) . There- 
fore as Fe 2+ concentration in solution 
increases, increasing quantities of base 
cations and protons will be desorbed from 
exchange sites because of competition from 
the Fe 2+ ions. For both Sphagnum species, 
when grown in solutions with Fe concentra- 
tions >^ 100 mg/L, not only did final Fe 
concentration in the plant tissues exceed 
the cation exchange capacity of Sphagnum 
(1 meq/g dry mass; cf. fig. 4), but also a 
dramatic reduction in both growth and 
chlorophyll concentration relative to 
controls was observed after 11 days (figs. 
1,2). Thus, the observed reduction in 
plant growth may have been a result of a 
deficiency of nutrient base cations 
induced by competition from Fe 2+ for 
exchange sites. 



Th 
tissues 
cation 
as much 
concent 
excess 
the pla 
cific b 
f ormati 
(cf. Wi 
oxide d 
examina 
scannin 
that or 
may hav 
Fe accu 
exchang 



e conce 
of bot 
exchang 
as 650 
rations 
Fe must 
nts as 
inding 
on of F 
eder et 
eposits 
tion of 
g elect 
ganic b 
e been 
mulat io 
e or ox 



ntrat 
h spe 
e cap 

% in 

1 10 

have 

nonex 

to or 

e oxi 

al. 

were 

the 

ron m 

indin 

a mor 

n tha 

ide f 



ion o 
cies 
acity 
Fe tr 
mg/ 

been 
chang 
ganic 
des o 
1987) 

disc 
tissu 
icros 
g of 
e imp 
n eit 
ormat 



f Fe 
excee 

of t 
eatme 
L (fi 

inco 
eable 

matt 
r oxy 
No 
ernib 
es us 
copy, 
Fe by 
ortan 
her c 
ion . 



in th 

ded t 
he pi 
nt so 
g. 4) 
rpora 

Fe v 
er or 
hydro 

obvi 
le du 
ing 1 

sugg 

the 
t pro 
ation 



he total 
ants by 
lution 

This 
ted into 
ia spe- 
by the 
xides 
ous 
ring 
ight or 
est ing 
plants 
cess in 



Sphagnum plants to Fe retention in a wet- 
land constructed for AMD treatment can be 
estimated based on the growth of the 
plants relative to controls (fig. 3) and 
measured Fe uptake (fig. 4). To estimate 
potential Fe uptake on an areal basis, 
field estimates of primary productivity of 
Sphagnum must be incorporated into the 
computation. The average annual net pri- 
mary productivity of the two dominant 
Sphagnum species from Big Run Bog ( S . 
f allax and Sj_ magellanicum ) is 5.75 g dry 
mass/dm 2 /yr (Wieder and Lang 1983) . Mult- 
iplying this value by the percent growth 
relative to controls (from fig. 3) and by 
final tissue Fe concentration (from fig. 
4) for each species at each solution Fe 
concentration yields estimates of poten- 
tial Fe accumulation by growing Sphagnum 
(table 1). The projected areal accumula- 
tion of Fe by S^ fallax was 30 to 230% 
greater than that by S_^ henryense for any 
particular solution Fe concentration. 
This difference is accounted for by the 
difference between the species in growth 
relative to controls; in all solution Fe 
concentrations excluding 10,000 mg/L, 
reduction of growth of S^ fallax was sig- 
nificantly less than the reduction of 
growth of S_^ henryense (fig. 3) . The 
maximum Fe accumulation projected for the 
two species (29 g/m 2 /yr) was more than 8 
times greater than a previous estimate 
(3.5 g/m 2 /yr) based on a lower Fe concen- 
tration in Sphagnum tissue (5.8 mg/g; 
Wieder et al. 1987). However, in achiev- 
ing an Fe uptake value of 29 g/m 2 /yr, the 
plants would be effectively dead. At Fe 
treatment solution concentrations at which 
plants were still growing and green, maxi- 
mum Fe uptake is only 6.5 g/m 2 /yr . 
Despite these new estimates of the poten- 
tial maximum for Fe uptake by Sphagnum 
plants, the contribution that plant uptake 
would make to total Fe retention in a 
wetland system is still small in compar- 
ison to the contributions made by other 
chemical and biological processes within 
the peat, especially the binding of Fe to 



Table 1. — Estimation of the potential Fe uptake, 
on an areal basis, by growing Sphagnum plants 
subjected to different Fe concentrations in 
AMD. Methods used to make these estimates 
are described in the text. 



The potential contribution of growing 



Treatment Solution 


Estimated 


Fe 


uptake 


(g/m 2 /yr) 


Fe Concentration 






















(mg/L) 


S. fallax 


£l 


henryense 


1 


2.2 








1.8 


10 


6.5 








4.1 


100 


16.8 








5.4 


1000 


29.0 








8.8 


10000 


17.7 








13.8 



283 



organic matter and the formation of Fe 
oxides, which together may account for as 
much as 97% of total Fe retention (Wieder 
et al. 1987) . 



Th 
plants 
with Fe 
at 100 
and sto 
to trea 
tively 
could r 
althoug 
more Fe 
trat ion 
Sphagnu 
compar i 
to Fe r 
for AMD 
Nonethe 
ous cov 
wetland 
terms o 
erosion 
demonst 
species 
respons 
greater 
henryen 
of Spha 



ese result 
can surviv 

concent ra 
mg/L or gr 
p growing, 
t low flow 
low Fe con 
emain viab 
h Sphagnum 

when expo 
s, the pot 
m plants t 
son to oth 
etention i 

treatment 
less, the 
er of vege 

treatment 
f minimizi 

of an org 
rated clea 

in their 
es to AMD; 

increase 
se in Fe-r 
gnum speci 



const it 
warrant 
Sphagnu 
exposed 



uents, par 

ed to dete 

m are most 

to differ 



s show tha 
e and grow 
tions up t 
eater, the 
In a wet 
s of AMD c 
centrat ion 
le. In ad 

plants do 
sed to hig 
ential con 
o overall 
er process 
n a wetlan 

appears t 
need to ma 
tation in 

systems i 
ng the pot 
anic subst 
r differen 
growth and 

S . f allax 
in length 
ich water, 
es respons 
ticularly 
rmine whic 

likely to 
ent types 



t Sphagnum 

in solutio 
o 10 mg/L, 
y turn brow 
land design 
ontaining r 
s, the plan 
dition, 

accumulate 
her Fe cone 
tribution o 
Fe retentio 
es contribu 
d construct 
o be minor, 
intain a vi 
constructed 
s critical 
ential for 
rate. We h 
ces between 
chlorophyl 
would exhi 
than S_^ 

Further s 
es to other 
to metals, 
h species o 
survive wh 
of AMD. 



ns 

but 

n 

ed 

ela- 

ts 



en- 
f 

n in 
ting 
ed 

gor- 

in 

ave 

1 
bit 

tudy 
AMD 
is 

f 
en 



ACKNOWLEDGMENTS 

Thanks to H. G. Spratt for his helpful 
comments on this manuscript, and to R. E. 
Andrus for Sphagnum i.d.'s. This work was 
supported in part by grants from the U. S. 
Environmental Protection Agency (R812379) 
and from the Sigma Xi Research Society, 
and by the villanova University Department 
of Biology. 

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Andrus, R.E. 1986. Some aspects of 

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416- 

Aulio, K. 

Spha 



426 

1 

gnu 



summ 
comp 
ombr 
Fenn 
Aul io , K . 
Spha 



an 
osi 
otr 
ici 
1 
gnu 



spec 
from 
habi 
101. 

Braekke, 
high 
Norw 
Fore 

Braunstei 
H.A. 



ifi 
om 

tat 



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neutral, treated coal mine drainage, 
pp. 433-441. I_n Symposium on surface 
mining, hydrology, sedimentation, and 
reclamation. University of Kentucky, 
Lexington, KY. 

Wieder, R.K., G.E. Lang, and A.E. White- 
house. 1985. Metal removal in 
Sphagnum -dominated wetlands: Experi- 
ence with a man-made wetland system, 
pp. 353-364 I_n Wetlands and water 
management on mined lands. J.E. 
Burris, D.E. Samuel, and J.B. Hill 
(Eds.). The Pennsylvania State Uni- 
versity, University Park, PA. 

Wieder, R.K., G.E. Lang, and A.E. White- 
house. 1987. Treatment of acid coal 
mine drainage using a Sphagnum - 
dominated wetland. pp. 218-220. In 
GEOMON- International Workshop on 
Geochemistry and Monitoring in Repre- 
sentative Basins. B. Moldan and T. 
Paces (Eds.). Prague, Czechoslovakia, 



285 



AN OVERVIEW OF THE ROLE OF ALGAE IN THE 
TREATMENT OF ACID MINE DRAINAGE 1 

D. A. Kepler 2 



Abstract. --Preliminary studies utilizing the digestion and 
ashing of freshwater algae of several phyla, samples from acid 
mine drainage (AMD) sources, indicate relatively high 
concentrations of iron and manganese associated with the algae. 
A wetland treatment system was constructed in September 1987 at 
an abandoned drift mine in Pennsylvania, associated with the 
Lower Clarion Coal Seam. In this system, specifically designed 
and planted "algae ponds" are integrated with emergent marsh 
sections of the system in an effort to examine the role of algae 
in treating AMD. The site of mineral accumulation in the algae 
(adsorbed and/or intracellular concentrations) and the form(s) 
in which the minerals are found in the algae are of primary 
interest. The relationship between iron and manganese regarding 
uptake in the cells, the life histories and possible succession 
of predominant algae species in the ponds, the correlation 
between mineral uptake and water quality, the effects of mineral 
toxicity to the algae, and the potential ability of algae to 
modulate effluent pH values on a diel basis will also be 
investigated. An overview of methods and their rationale along 
with results to date at the noted study site will be presented, 
with the intent of allowing for the replication and expansion of 
the study by others. 



INTRODUCTION 

Freshwater algae are common and widespread 
inhabitants of the waters associated with coal mine 
drainage and are recognized as being effective 
accumulators of metals (Bates et al. 1982, Crist 
et al. 1981, Wetzel 1975). The majority of work 
related to algae and the bioaccumulation of metals 
has typically been with metals that are generally 
of little concern, from a regulatory standpoint, in 
coal mine drainage, notably zinc, nickel, copper, 
and lead. 



'Paper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the 
Interior (Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), 
April 17-22, 1988, Pittsburgh, PA. 

2 D. A. Kepler, Environmental Specialist, The 
EADS Group, General Engineering Division, 
Clarion, PA. 



Iron and manganese, however, are probably the 
two most common, and regulated, mineral 
constituents of mine drainage, accounting for 
millions of dollars in chemical treatment costs 
annually. Iron, which is relatively easy to 
remove from mine drainage, presents few logistical 
problems to chemical treatment, but manganese is a 
more persistent mineral and requires rather 
sophisticated treatment to achieve effluent 
standards. Additionally, concentrations of 
50-100 mg/L of manganese are not uncommon in acidic 
mine drainage (Patterson 1975). 

Algae require iron and manganese as essential 
micronutrients . Although concentrations of 
manganese > 1 mg/L have been shown to be inhibitory 
to some blue-green and green algae (Gerloff and 
Skoog 1957, Patrick et al . 1969), algae are 
commonly found and often abundant in AMD with high 
concentrations of iron and manganese. Algae 
associated with these areas have been shown to 
accumulate manganese, with reported levels of 
manganese as high as 56,000 mg/kg of plant tissue 



286 



(dry weight) (Kepler 1986). This is of particular 
interest in that biological treatment systems, 
i.e., man-made wetlands, are proving to be a viable 
and cost-effective means of treating AMD. 

To date, most man-made wetlands are 
constructed as Sphagnum -dominated , or more 
commonly, Typha -dominated treatment systems 
(Kleinmann and Girts 1986). Algae are generally 
not purposefully introduced into these wetlands, 
but become established in the systems by way of the 
source of the mine drainage, the air, soil, and 
most commonly, in association with the principal, 
transplanted wetland vegetation. Because of this 
secondary nature of establishment, any beneficial 
effects of algae on the quality of AMD often go 
unrecognized, or the beneficial effects are limited 
because the Wetland Treatment System (WTS) is not 
designed to optimize algal habitat and therefore, 
algal bicmass is limited. Little is actually known 
concerning the role algae has in the treatment of 
AMD in these Wetland Treatment Systems. 

This paper is intended to be an overview of 
work in progress concerning the role of algae in 
the treatment of AMD in WTS's, with particular 
emphasis placed on the mechanisms involved in the 
bioaccumulation of manganese, Construction of the 
study WTS began in September of 1987, and 
therefore, limited findings are available at this 
time. However, the objectives and methods of the 
study, along with any results to date, will be 
reported and discussed. 



DESCRIPTION OF RESEARCH 

Background 

This study is being funded by the United 
States Department of the Interior, Bureau of Mines, 
under the Abandoned Mine Land research program and 
is listed as, "Acid Mine Drainage Treatment 
Utilizing Blue-Green Algae." The study wetland is 
located in Richland Township, Venango County, PA, 
and was constructed in cooperation with Glacial 
Minerals, Inc., of Strattenville, PA, on land owned 
and previously surface mined by Glacial Minerals, 
Inc. The source of the acid mine drainage being 
treated by the WTS is from an abandoned drift mine 
(circa 1920) that has been discharging since at 
least 1940 and is now considered to be receiving 
recharge waters from the adjoining surface mine. 
The water is associated with the Lower Clarion Coal 
Seam and flows at an average of 37 gal/mi n. The 
AMD quality at the source is characterized as 
follows: pH, 5.0; alkalinity, 16 mg/L; mineral 
acidity, 450 mg/L; total iron, 130 mg/L; manganese, 
71 mg/L; aluminum, 4 mg/L; and sulfate, 2,200 mg/L. 



Construction 

Construction began in September of 1987 with 
the majority of the system completed by November of 
that year. The WTS was designed and built with an 
initial collection pond immediately adjacent to the 
drift mine source, followed by two, distinct, 
parallel wetlands. The two wetlands combine at 
their respective outflow points and flow through a 
planted ditch to a conventional treatment pond 
series . 



The parallel wetlands are set up as four 
individual units in each system, two shallow 
(6-12 in), Typha -dominated ponds followed by two, 
deeper (3-4 ft) algae-dominated ponds in one 
series; and two algae ponds followed by two Typha 
ponds in the other series. Each pond has an 
approximate surface area of 600 ft 2 and is 
connected to the next pond in series by a section 
of 6-in diameter plastic pipe. The system is 
graphically displayed in figure 1. 

The ponds were constructed with available, on- 
site clay and a 12-in base of composted manure 
which serves as a planting medium and a source of 
nutrients to the vegetation. Four to six inches 
of agricultural limestone was placed beneath the 
compost in the final, planted ditch, but no 
alkaline materials were used in any other portion 
of the wetland. Typha latifola was available on 
site, and was hand dug and planted at a density of 
roughly one mature plant and rhizome mass per 
2 ft 2 . Scirpus cyperinus and J uncus effusus were 
also available on site and were introduced in the 
Typha ponds primarily in the areas of the 
inflow/outflow pipes. S cyperinus and J^ effusus 
were also planted around the edges of the algae 
ponds to stabilize the banks and to provide 
possible sites of attachment for algae. 

Initially, the study intended to concentrate 
on blue-green algae (Cyanobacteria) of the genus 
Oscillatoria . Oscillatoria spp. was noted as being 
prevalent in several natural and man-made wetlands 
that displayed reductions in manganese levels from 
source to outflow, and was the species noted above 
exhibiting manganese levels of 56,000 mg/kg of 
plant. However, during the period between the 
original proposal and the actual planting of the 
study wetland, the predominant algae in the 
previously Oscillatoria - dominated wetlands became 
the green algae of the genus Mougeotia . 
Oscillatoria was still present, but not dominant. 

The water quality in these wetlands continued 
to show manganese reductions, and analysis of 
Mougeotia samples confirmed that they too were 
accumulating manganese. Therefore, the decision 
was made to transplant algae from one of these 
sources into the study WTS, knowing that the 
predominant algae being "planted" was Mougeotia , 
but that Oscillatoria was also being introduced at 
the study site. Approximately one 30-gal container 
of Mougeotia - dominated algae was introduced into 
each of the four algae ponds at the study wetland. 
Because the side of the parallel series with the 
Typha ponds down-flow of the algae ponds would be 
immediately affected by this planting and the other 
side would not be affected, a container of algae 
was also introduced into each of the four Typha 
ponds . 

Various other species of algae, primarily 
diatoms, were also introduced into the wetland with 
the principal algae planting, and a filamentous 
green algae was noted as being associated with the 
Typha planting. These species, while of little 
apparent significance regarding total numbers at 
the time of the planting, are being identified and 
may be found to take on some degree of importance 
in treatment during the 1988 growing season. 



287 



Figure 1 . — Plan view of wetland treatment system. 




2a 



■ 



7 
> 



■ 



1 1 






m 

mm 



&m 





1 

ft 
1 



&^M%M&%^M^^^ 



Nos. = mon. pts. 



^5v 



= Typha 



algae ponds 



Water Analysis 

The AMD source, the inflow/outflow for each 
pond, and the final outflow point of the wetland 
have been designated as water monitoring points in 
this study. Each monitoring point will be sampled 
weekly over the 1 8-month study which began in 
September of 1987. The samples are analyzed for pH 
(electrometrically) ; specific conductance (p mhos 
at 25° C); alkalinity and mineral acidity 
(titrimetrically at pH 4.5 and 8.3 respectively); 
and total and ferrous iron, total and dissolved 
(dissolved: passing a M5-y membrane filter) 
manganese, and sulfate and aluminum (all minerals 
are determined spectrcphotometrically) . Water 
temperatures (° C) are taken from each pond in the 
field at the time of collection. All analyses are 
conducted following the Standard Methods for the 
Examination of Water and Wastewater and are done in 
accordance with E.P.A. guidelines. 

Flows are measured weekly by means of a 
portable cutthroat flume in the channels and by 
bucket and stopwatch at the discharge pipes. In 
this manner, the total water budget of the WTS may 
be known, with any increases or decreases in the 
system readily noted. 



Vegetation Analyses 



Representative samples 
vegetation will be collected 
manganese, and aluminum cont 
spring of 1988 and then peri 
the study, with principal co 
predominant algae species 
determined as follows: the s 
ashed by means of heating at 
furnace for two hours, acid 
methods, and analyzed for mi 
atomic absorption. Results 
mineral/kg of plant (dry Wei 



of the wetland 

and analyzed for iron, 
ent beginning in the 
odically throughout 
ncern given to the 
Total metals are 
amples are dried, 

550° C in a muffle 
digested by standard 
neral concentrations by 
are reported as mg of 
ght). 



To determine the site of mineral accumulation 
in the algae, i.e., adsorbed and intracellular 
concentrations, the algae samples are "washed" with 
an EDTA solution in a shaker bath and then 
filtered. The extract is analyzed by atomic 



absorption to determine the concentration of 
adsorbed mineral associated with the algae. The 
"washed" algae is then ashed and analyzed as 
described above to determine the intracellular 
concentration of minerals. 

The form of the minerals in and on the algae, 
e.g., adsorbed manganic oxide, will also be 
determined utilizing a form of x-ray diffraction so 
as to predict the long-term stability of the 
minerals in the wetland and the mechanisms of 
accumulation. 

Diurnal studies will be conducted during 
periods of accelerated algal growth to examine the 
dissolved oxygen content and pH of the ponds and, 
therefore, whether the algal blooms have any 
effect on pH values. Carbon dioxide removal and 
production in the wetland, through algal 
photosynthesis and respiration, can account for 
shifts in pH values, although this phenomenon is 
generally associated with alkaline waters (Cole 
1979). Lake pH's in western Pennsylvania have 
been noted to fluctuate five pH units over a 24-hr 
period due to the effects of blue-green algal 
blooms (Kepler 1985). 



RESULTS AND DISCUSSION 

Water 

At the time of this writing, water quality 
data are available for the last two weeks of 
September 1987, through the first two weeks of 
January 1988. The numbers reflect the effects of a 
newly planted, somewhat dormant wetland. The Typha 
planting died back early in September simply 
because of the season, and the algae populations 
have not had favorable weather conditions to bloom. 
However, between the compost and the dead 
vegetation there is considerable organic matter 
present in the WTS, the rhizomes in the Typha 
ponds are active, bacteria and algae are present 
in the system, and the WTS is designed to maximize 
plant contact and retention time with the flows. 
Therefore, even at this early date, positive water 
quality results can be seen (figures 2-5). 



288 



200 




9/87 10/87 11/87 12/87 

Date 
Figure 2. --Average total iron at AMD source, 
points 6 & 10, and final outflow. 



The WTS is significantly reducing the iron 
and acidity load of the AMD, but manganese values 
are not significantly reduced at this time. To 
date, the quantity of water entering and exiting 
the system has been equal. A small amount of 
seepage has been noted along the coal outcrop, 
upslope and immediately adjacent to the pond 
series ending with the Typha ponds (point 6), but 
the seepage is small enough that neither the 
quantity or quality of the flows in the WTS has 
been affected. The outcrop area has been planted 
in an attempt to control the seepage, since any 
drainage entering the WTS near its middle or end 
would obviously bias any results or conclusions 
regarding the effectiveness of the WTS. 



Vegetation 




■ Point 1 

Point 6 

M Point 10 

Point 11 



9/87 10/87 11/87 12/87 

Date 

Figure 3.-Average total manganese at AMD 
source, points 6 & 10, and final outflow. 



80 




Point 1 
Point 6 
Point 10 
Point 1 1 



9/87 10/87 11/87 12/87 
Date 

Figure 4. -Average dissolved manganese at AMD 
source, points 6 & 10, and final outflow. 



500 




I Point 1 

Point 6 

^ Point 10 

Point 11 



9/87 10/87 11/87 12/87 

Date 
Figure 5. -Average mineral acidity at AMD source, 
points 6 & 10, and final outflow. 



The actual vegetative analyses will not begin 
until spring 1988 unless the algae ponds bloom 
during the winter months. The operational 
definition for "adsorbed minerals," in regard to 
the algae, is "that portion of the total minerals 
which can be extracted with the EDTA wash." The 
washing and filtration of the samples undoubtedly 
stresses the algae and may therefore bias the 
findings in relation to the concentration of 
adsorbed minerals. However, a similar study 
utilizing an EDTA wash to extract adsorbed zinc 
from algal samples indicated that the technique 
had no appreciable effect on the total 
concentration of surface-bound zinc, in that 
measurable intercellular zinc was not released by 
the technique (Bates, et al . 1982). 

The most efficient method of mineral 
extraction for this study will be determined during 
the noted winter refinement period by testing 
varying lengths of extraction time, EDTA 
concentrations, and method of supernatant 
collection. The definition of adsorbed minerals 
will remain operational, however, since it will be 
impossible to unequivocally state that no 
intercellular minerals were released, and 
therefore measured, during the extraction process. 

The ability to distinguish adsorbed from 
intracellular minerals will allow for a more 
complete understanding of the role of algae and 
the mechanisms of mineral accumulation in WTS's. 
Algal cell walls are composed of numerous, 
potential mineral adsorption sites as a result of 
their chemical makeup; however, these sites may or 
may not be active in transporting iron and 
manganese within the cell. The cell walls can act 
as both chemical and physical barriers in 
regulating mineral transport within the cell, 
thereby providing a means of metal tolerance to the 
algae by excluding potentially toxic ions from 
entering the cell (Foster 1977). 

It is likely that the ashing data will show 
the majority of iron and manganese will be 
adsorbed versus intracellular; this is chiefly 
because of the known large concentrations of these 
minerals associated with the algae in the 
background ashings and the minerals' potential 
toxicity at these levels to the living cells. 

Appropriate statistical analyses will be 
employed to determine whether intracellular 
concentrations of minerals are a function of the 
adsorbed concentrations, or if they are transported 



289 



within the cell independently of the surface 
concentrations. Previous data also suggest that 
there may be an inverse relationship between iron 
and manganese accumulation in wetland vegetation 
(Kepler 1986), a relationship that will be tested 
in this study. The data will be further analyzed 
to determine if there is a correlation between 
total mineral concentrations in the AMD and total 
mineral uptake and accumulation in the algae; 
i.e., whether algae are more effective 
bioaccumulators at particular AMD mineral 
concentrations. Add to these findings the benefit 
of knowing the chemical form of the accumulated 
minerals, and predictions as to the success and 
long-term stability of WTS's will become more 
accurate. 



Patrick, Ft. , B. Crumb, and J. Coles. 1969. 

Temperature and manganese as determining 
factors in the presence of diatom or blue- 
green algal floras in streams. Proc. Nat. 
Acad. Sci. 61:172-178 

Patterson, J. W. (Ed.). 1975. Wastewater 

treatment technology. 265 pp. Ann Arbor 
Science, Ann Arbor, MI. 

Wetzel, Ft. G. (Ed.). 1975. Limnology. 713 pp. 

Saunders College Publishing, Philadelphia, PA. 



SUMMARY 

The removal of manganese in WTS's integrated 
with specific algae "ponds" may prove to be a 
successful treatment technique. Algae appear to be 
particularly well suited to this type of system in 
that they are widespread in AMD, are capable of 
accumulating relatively high concentrations of 
minerals, and are capable of rapid reproduction. 
The success of this single WTS and the mechanisms 
of the bioaccumulation of AMD minerals in algae 
will be determined in the proceeding months. 



LITERATURE CITED 

Bates, S. S. , A. Tessler, G. C. Campbell, and 
J. Buffle. 1982. Zinc adsorption and 
transport by Chlamydomonas variabilis and 
Scenedesmus subspicatus (Chlorophyceae) grown 
in semi continuous culture. J. Phycol. 
18:521-529 

Cole, G. A. (Ed.). 1979. Textbook of Limnology. 
126 pp. C. V. Mosby Company, St. Louis, MO. 

Crist, R. H. , K. Oberhosler, N. Shank, and 

R. Nguyen. 1981. Nature of bonding between 
metallic ions and algal cell walls. Environ. 
Sci. Technol. 15:1212-7. 

Fester, P. L. 1977. Copper exclusion as a 

mechanism of heavy metal tolerance in green 
alga. Nature (Lond.) 269:322-3. 

Gerloff, G. C, and F. Skoog. 1957. Availability 
of iron and manganese in southern Wisconsin 
lakes for the growth of Microcystis 
aeruginosa . Ecology. 38 : 551 -556. 

Kepler, D. A. 1985. Unpublished data from 

observations of natural and man-made wetlands 
in western Pennsylvania. 

. 1986. Manganese removal from mine 



drainage by artificial wetland construction. 
In: Proceedings of the 8th annual national 
abandoned mine lands conference, pp. 71-80. 
Billings, MT. 

Kleinmann, R. L. P., and M. A. Girts. 1986. 

Constructed wetlands for treatment of mine 
watei — successes and failures. In: 
Proceedings of the 8th annual abandoned mine 
lands conference, pp. 67-73- Billings, MT. 



290 



TRACE METAL REMOVAL FROM STOCKPILE DRAINAGE BY PEAT 1 



Kim Lapakko and Paul Eger 



Abstract. — Batch reactor tests were conducted 
to determine the ability of peat to remove trace 
metals (Cu, Ni, Co, Zn ) from one acidic and two near 
neutral mining stockpile drainages. In all three 
drainages, nickel contributed 70 to 85 percent of 
the trace metal concentration. The sum of the trace 
metal concentrations in the acidic drainage was 580 
mg/L at pH 3.15. At a loading of 50 g dry peat/L, 
60 percent of the trace metals were removed from 
this drainage after one hour of reaction, with no 
subsequent removal. The trace metal concentrations 
in the near neutral drainages totaled 1.2 mg/L at pH 
7.4 and 9.2 mg/L at pH 7.9. Kinetic studies on 
these two drainages indicated that the majority of 
metal removal occurred in the first 20 hours of 
reaction and that equilibrium was approached after 
about 70 hours. At a loading of 5 g dry peat/L, 
metal removal of SO to 90 percent was observed for 
solutions maintained at pH 7.4. Trace metal removal 
decreased as pH decreased from 7.4 to 4.0. Release 
of the bound metals to distilled water at pH 7.4 was 
slight, and was not greatly affected by elevated 
concentrations of calcium, magnesium, sodium, 
potassium and sulfate. Metal release increased as 
solution pH decreased and was greater yet when the 
peat was exposed to bog water. Column experiments 
were conducted on the near neutral drainage. Metal 
removal from the acidic drainage was 19 mg trace 
metal/g dry peat. 



Paper presented at the 1988 Mine Drainage and Surface Mine Reclamation Conference 
sponsored by the U.S. Department of the Interior (Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), April 17-22, 1988, Pittsburgh, PA. 

2 

Kim Lapakko is Senior Engineer and Paul Eger is Principal Engineer, Minnesota Department 

of Natural Resources, Division of Minerals, St. Paul, MN. 



291 



INTRODUCTION 

Five stockpiles containing low levels 
of metal sulfides are located at the LTV 
Steel Mining Company Dunka Site. Nickel 
concentrations in the drainages from these 
piles have ranged from less than 1 mg/L to 
nearly 40 mg/L. Concentrations of copper, 
cobalt, and zinc are also elevated. The 
Minnesota Pollution Control Agency (MPCA) 
recently concluded that metal concentra- 
tions in the creek draining the mine site 
were excessive, and in 1986 required that 
LTV examine means of reducing metal 
release from the watershed. The monthly 
average total concentration goals for in 
stream copper, nickel, cobalt and zinc 
were 0.030, 0.130, 0.010, and 0.047 mg/L, 
respectively. These values were based on 
United States Environmental Protection 
Agency (EPA) water quality criteria and 
other aquatic toxicity data (EPA 1980, 
1984, 1986; MPCA 1988). The initial 
feasibility study indicated that wetland 
treatment may be a viable mitigation 
techniques for drainages with low or high 
trace metal concentrations. 

In recent years interest has risen in 
the use of natural and constructed wet- 
lands for treatment of acid mine drainage. 
Such applications are of particular 
interest in Minnesota, where peatlands 
cover an area of roughly 30,000 km . The 
focus of coal mine drainage treatment has 
been on increasing pH while reducing 
concentrations of acidity and iron. As 
mine drainage flows through wetlands, 
trace metals can also be removed from 
solution due to sequestration by peat, 
vegetative uptake, precipitation of metals 
with sulfide produced by sulfate reduc- 
tion, and/or precipitation as metal 
hydroxides under conditions of elevated 
pH. 



The occurrence of ele 
metal concentrations in pe 
natural conditions, has lo 
nized by geologists as an 
mineralization (Fraser 196 
1962 it was suggested that 
a medium for removing meta 
waters (Oreshko et al. 196 
geochemical enrichment fac 
of the metal concentration 
concentration in the solut 
the peat is in contact) fo 
and Zn have been reported 
6,900 and 8,600, respectiv 
and Khan 1972). Copper co 
2.2 to 8.9 percent of the 
have been reported, as hav 
zinc concentrations of 2.3 
cent, respectively (Premi 
al. 1977, Szabo 1958). 



vated trace 
at , under 
ng been recog- 
indicator of 
1). As early as 

peat be used as 
Is from waste 
2). Approximate 
tors (the ratio 

in peat to its 
ion with which 
r Cu, Ni, Co, 
as 2,400; 450; 
ely (Schnitzer 
ncentrations of 
dry peat mass 
e nickel and 

and 3 . 3 per- 
1970, Smith et 



The following conclusions were 
extracted from a literature review which 
addressed metal sequestration mechanisms 
and capacities of peat, as well as appli- 
cation of peat to waste water treatment 
(Lapakko et al. 1986). Ong and Swanson 
(1966) cited physical adsorption as the 



force which bound copper to peat. However, 
the majority of researchers concluded that 
the metal-peat bond is chemical in nature, 
specifying chemisorption , complexation , or 
chelation. It is not unlikely that both 
physical and chemical forces are involved 
in the reaction. Several researchers 
reported that ion exchange is involved in 
the reaction. To examine the sequestration 
of trace metals from mine drainage by peat, 
batch reactor and column studies were 
conducted. The laboratory results were 
compared to values observed in a white 
cedar swamp through which stockpile drain- 
age flowed naturally (Eger et al. 1980). 



METHODS 

The peat used was a fibric, woody peat 
collected from the upper 0.75 m of an 
unimpacted black spruce bog in northeastern 
Minnesota. The pH of the peat was 4.5, as 
determined for a mixture of 15 cm dry peat 
and 15 mL water, and the cation exchange 
capacity determined by the ammonium acetate 
method was 105 meq/100 g. Three drainages 
were examined, one of which was acidic 
(table 1). Drainage B was generated by a 
waste rock (0.057 pet Cu , 0.014 pet Ni) 
stockpile at the LTV Dunka site. The other 
two drainages were generated from small 
scale test stockpiles located near the 
Dunka site. The pile (FL1) generating 
Drainage A contained 0.35 pet Cu, 0.083 pet 
Ni, and 0.63 pet S. The pile (FL5) gener- 
ating Drainage C contained 0.34 pet Cu, 
0.084 pet Ni, and 1.41 pet S. 

Erlenmeyer flasks (250 mL) covered 
with parafilm were used as batch reactors. 
A well-mixed condition was maintained by 
agitating the reactors on an Eberbach 6150 
rotating shaker which was set at 125 
excursions per minute. Kinetic experiments 
were conducted with 100-mL volumes of 
Drainages A, B, and C using peat loadings 
of 0.5, 1.0, and 50 g/L, respectively. 
Only one additional experiment was con- 
ducted with the acidic Drainage C. Metal 
removal was determined at seven peat 
loadings ranging from 0.5 to 50 g/L; 
solution pH was not controlled. 

Experiments with Drainage A and 
Drainage B were more extensive. In batch 
reactor tests, metal removal at pH 7.4 was 
examined using peat loadings of 0.1 to 7 
g/L. Three additional pH values, 4.0, 5.0, 
and 6.0, were also examined at three 
different mass loadings. pH was maintained 
by daily manual addition of NaOH or H-SO. 
at concentrations of 1, 0.1, or 0.01 N, 
depending on the magnitude of adjustment 
required. Following the trace metal 
removal phase, release of metals from the 
solids was examined in a rinse phase with 
distilled water, a salt solution, and bog 
water. The distilled water rinse solutions 
were maintained at the same pH as the 
drainage solutions in the removal phase. 
The ions in the salt solution were Ca, Mg, 
Na, K and SO.. 



292 



Table 1. — Composition of Drainages. 

(Concentrations in mg/L unless otherwise noted. ) 



Batch drainages 
B 



Column influent 
A B 



Cedar, 
swamp" 



Cu 

Ni 

Co 

Zn 

CEU 

Fe 

Mn 

Ca 

Mg 

Na 

K c 

TAlk* 

TOC 



7.4 


7.9 


3.2 


7.2-7.4 


7.2-7.9° 


7.2 


0.046 


0.08 


110 


0.03 


0.02 


0.62 


1.06 


8.6 


430 


.86 


5.4 


18 


.029 


.33 


26 


.05 


.17 


1.2 


.10 


.22 


16 


.11 


.17 


.38 


.19 


1.3 


180 


.18 


.75 


3.7 


NA 


NA 


NA 


NA 


NA 


1.4 


NA 


NA 


NA 


NA 


NA 


6.5 


510 


300 


390 


270 


230 


280 


62 


142 


520 


26 


130 


220 


150 


98 


210 


80 


90 


NA 


50 


10 


34 


30 


8 


NA 


NA 


NA 


NA 


18 


51 


110 
22 


NA 


NA 


NA 


10 


16 



NA: not analyzed. 

^Arithmetic mean value of stockpile drainage input over 14 months. 
„In standard pH units 

.Influent pH values varied over the course of the 2 to 4 month experiment. 
jXEU = [Cu] + [Co] +0.1 ([Ni] + [Zn]), with concentrations in mg/L. 
fi Total alkalinity as CaC0„ 
-Total organic carbon as C 
Dissolved organic carbon. 



Column experiments were conducted 
with these two drainages. A 2:1 mixture 
of wet peat to sand was required to 
maintain adequate flow rates in columns 
designed for a saturated flow rate of 5 
bv/day (bed volumes per day, where one bed 
volume is the space occupied by the treat- 
ment bed, including voids). A saturated 
hydraulic conductivity of 5.4 x 10 
cm/sec was determined for the peat-sand 
mixture at a dry bulk density of 1.2 
g/cm . The peat-sand mixture in columns,, 
had a dry bulk density of 1.0 to 1.2 g/cm 
and a porosity of 0.4 to 0.5. 

Aqueous samples were filtered (0.45 
micron) and analyzed for pE , sulfate, 
trace metals (Cu, Ni, Co, Zn), major 
cations (Ca, Mg , Na, K) , and specific 
conductance. Solution pE and specific 
conductance were analyzed in the labora- 
tory using an Orion 601A pE meter with a 
9104 electrode and a Myron L conductivity 
meter, respectively. Serco Laboratories 
analyzed for sulfate using the barium 
sulfate turbidimetric technique. Metals 
were analyzed by atomic absorption in 
either the flame or flameless mode at the 
Minnesota Department of Natural Resources 
Minerals Division Office in Bibbing, MN. 



RESULTS 

To simplify presentation of results, 
the total trace metal concentration is 
occasionally presented in terms of "copper 



equivalent units" or CEU, which is calcu- 
lated as: 

CEU = [Cu] + [Co] + 0.1([Ni] + [Zn]) (1) 

This notion is based on data indicating 
that nickel and zinc are only one tenth as 
toxic as copper and cobalt (Lind et al. 
1978). 

A preliminary experiment was con- 
ducted to select mass loadings at which 
kinetic experiments would be conducted. 
This experiment indicated that as the peat 
loading increased, the solution pE 
decreased. For Drainage A and Drainage B 
the pE was reduced from 7.4 and 7.9, 
respectively, to about 3.7 at a peat 
loading of 50 g/L (dry weight). To limit 
the influence of peat on pE , and a poten- 
tial resultant influence of pE on the 
kinetics of metal removal, loadings of 5 
and 10 g/L were selected for Drainages A 
and B, respectively. The peat-induced 
reduction of Drainage C pE from an initial 
value of 3.22 was less than 0.3 units, and 
a loading of 50 g/L was chosen for the 
kinetic experiment with this drainage. 

Solution pE was not controlled in the 
kinetic experiment, and in all cases it 
decreased over time. The initial pE of 
Drainage C was 3.1, dropping during 
storage from the value in the loading 
experiment. The pE decreased only 
slightly to 3.03 over 168 hours. For each 
trace metal, removal occurred within the 
first hour (fig. 1). With Drainages A and 



293 



140 




LEACHATE C KINETICS 








120 










- 


100 












SO 








ni 
















60 










- 


40 




















G Cu 




20 


















i i i 


1 


1 




TIME (HOURS) 



TIME (HOURS) 



Figure 1. — Aqueous trace metal 

concentrations in Drainage C with a 
50 g/L peat loading as a function of 
time . 



Figure 2. Aqueous trace metal 

concentrations in Drainage B with a 
g/L peat loading as a function of 
time . 



B the pH was more affected and trace metal 
removal was less rapid. The pH decreased 
from 6.9 to 5.9 with Drainage A and from 
7.7 to 6.1 with Drainage B. The majority 
of trace metal removal from these two 
drainages occurred within 24 hours and 
little removal occurred after 48 hours 
(fig. 2). 

Isotherm experiments, in which the 
peat loading was varied, were conducted on 
each of the three drainages. Since the pH 
of Drainage C did not vary greatly with 
peat loading, data from the preliminary 
loading experiment were used. Peat 
loadings of 0.5, 1.0, 5.0, 10, and 50 g/L 
were used in this experiment. Equilibrium 
pH values fell from 3.22 for the untreated 
Drainage C down to 2.95 at a peat loading 
of 50 g/L. There was no detectable trace 
metal removal at peat loadings of 0.5 and 
1.0 g/L and less than 20 percent removal 
at 10 g/L. At the maximum loading of 50 
g/L, copper removal was about 80 percent 
while removal of nickel, cobalt, and zinc 
was all about 40 percent. Mass removal of 
nickel was the greatest of the metals, 
with concentrations reduced from 430 to 
280 mg/L (fig. 3). No further experiments 
were conducted with Drainage C. 



Experimentation with Dr 
B was more extensive. To el 
effects on isotherm experime 
drainages, the pH in the rea 
maintained at 7.4 by regular 
acid or base. With Drainage 
ings of 0.1, 0.2, 0.5, 1.0, 
5.0 g/L were used. Based on 
and mass reduction, nickel r 
greatest. From an initial v 
mg/L, nickel concentrations 



ainages A and 
iminate pH 
nts with these 
ctors was 

addition of 
A, peat load- 
2.0, 3.0, and 

both percent 
emoval was the 
alue of 1.06 
strictly 




20 30 40 
PEAT LOADING, G/L 



Figure 3. — Trace metal concentrations in 
Drainage C as a function of peat 
loading. 



decreased to 0.78 and 0.09 as the peat 
loading increased from 0.1 to 5.0 g/L. The 
copper concentration of 0.046 mg/L was 
reduced to 0.013 mg/L at the 0.1 g/L 
loading and typically remained in the range 
of 0.007 to 0.009 mg/L over the remaining 
loadings. Cobalt and zinc concentrations 
were fairly constant with respect to 



294 



Z 4 

o 



r- 


i ! 


1 —'T 1— 


— i 


) \ 






- 


>\ 






- 


-\ 


\ Co 




- 


\ni 














- 


} N 






- 






^H—CY^^^ 


- 


Cu 








J-C^-G)— 


— (? nv 


— -t m i— 





10.0 



.25 Z 

O 



o 

10 J 



3 4 5 

PEAT LOADING, G/L 



Figure 4. — Trace metal concentrations in 
Drainage B as a function of peat 
loading. 



loading, with equilibrium values between 
0.04 and 0.06 mg/L. There was an apparent 
release of cobalt to solution, but this 
was most likely attributable to an error 
in the analysis of the untreated drainage. 

With Drainage B, peat loadings of 
0.1, 0.5, 1, 2, 3, 5, and 7 g/L were used. 
From an initial value of 8.6 mg/L, nickel 
concentration decreased to 6.0 at the 0.1 
g/L loading. Concentrations continued to 
decrease as loading increased and reached 
a minimum of 0.26 mg/L at the 7 g/L 
loading, a 97 percent concentration 
reduction. The reduction in copper 
concentration exceeded 90 percent at all 
loadings, with equilibrium concentrations 
in the range of 0.005 to 0.007 mg/L. The 
cobalt and zinc concentrations tended to 
decrease as the peat loading increased, 
with a maximum reduction of about 80 
percent. The minimum observed aqueous 
concentrations of cobalt and zinc were 
0.06 and 0.04 mg/L, respectively (fig. 4). 
Concentrations of calcium, magnesium, 
sodium, and potassium were also analyzed 
in this experiment. Calcium was the only 
one of these parameters for which removal 
was observed. Concentrations were reduced 
from 300 to 180 mg/L at the 70 g/L peat 
loading. Concentrations of magnesium and 
potassium were quite stable, typically 
remaining within ten percent of the 
initial values. Sodium concentrations 
increased only slightly, with an observed 
increase of 20 mg/L at the 7 g/L peat 
loading. 




pH 



Figure 5. — Trace metal concentrations in 
Drainage B as a function of pH at a 
peat loading of 7 g/L. 



The influence of pH on trace metal 
removal was examined in batch reactor 
experiments with Drainages A and B. With 
each drainage, three different peat load- 
ings were exposed to drainage maintained at 
pH values of 4, 5, 6, and 7.4. In all 
cases trace metal removal decreased with 
decreasing pH (fig. 5). The results for 
equilibrium CEU levels were quantified as 
follows : 



log CEU e = a (pH) 



+ b 



(2) 



The constants determined by linear regres- 
sion are presented in table 2. 



To determine 
metals were bound 
in the isotherm t 
B at pH 7.4 was p 
mL of distilled w 
the isotherm test 
was maintained at 
of acid or base, 
to distilled wate 
typically less th 
sorbed metal. 



how strongly the sorbed 
each of the solids used 
ests with Drainages A and 
laced in contact with 100 
ater and agitated as in 
s. The pH of the solution 
7.4 by regular addition 
Release from the solids 
r at pH 7.4 was small, 
an five percent of the 



295 



Table 2.-- Linear regression constants for log CEU = a pH + b" 



Peat 
Load 
(g/L) 



Drainage A 



Peat 
Load 
(g/D 



Drainage B 



0.1 

1 
5 



-0.038 

- .13 

- .17 



-0.532 

- .127 

- .15 



0.74 
.98 
.95 



-0.047 

- .19 

- .28 



0.38 

.66 

1.11 



0.64 
.59 
.98 



CEU represents the equilibrium aqueous CEU concentration in mg/L. 



Rele 
of the ri 
ionic str 
with Drai 
2.5 to 5 
a salt so 
observed 
pH. Thes 
solids us 
drainages 
pH 4 was 
distilled 



ase tended to 
nse solution 
ength increas 
nage B, metal 
times that at 
lution at pH 
for distilled 
e trends were 
ed with Drain 

metal rele 
six to eight 
water at the 



increase as the pH 
decreased and as its 
ed. For solids used 
release at pH 4 was 
pH 7.4. Release to 
7.4 was twice that 
water at the same 
less pronounced for 
age A. For both 
se to bog water at 
times the release to 
same pH. 



Trace metal removal from Drainage A 
and B was also examined using a set of 
triplicate columns for each drainage. The 
results from each set of three columns 
were highly reproducible. The peat-sand 
mixture used in column tests with Drainage 
A depressed pH. Effluent values of 3.9 to 
4.1 were observed, as compared to the 
influent pH of 7.4. With this drainage 
the peat-sand was replaced with a mixture 
of peat and low sulfur tailings (2 g wet 
peat: 1 g dry tailings) which produced a 
higher effluent pH and lower trace metal 
concentrations than the peat-sand mixture. 

The columns containing the peat- 
tailings mixture reached breakthrough 
after 230 bed volumes (bv) of Drainage A 
flowed through the column at a rate of 5 
bv/day. The maximum CEU reduction ob- 
served was 50 percent. The removal of 
individual metals tended to increase with 
the influent concentration (fig. 6). 
Removal was lowest for cobalt (essentially 
zero), and copper (6.3 mg/kg dry solid), 
and highest for nickel and zinc (85 and 11 
mg/kg dry solid, respectively). 

Although the Drainage B columns did 
not reach total breakthrough (the point at 
which concentrations in the effluent equal 
those in the influent), the trends were 
similar to those observed for the Drainage 
A columns and in batch experiments. Metal 
removal was a function of trace metal 
concentration and increased as pH in- 
creased. The initial effluent had ele- 
vated trace metal concentrations, although 
less than influent levels, and depressed 
pH. The trace metal concentrations 
decreased fairly rapidly, reaching a 
minimum after about 25 to 35 bv of flow. 
Concurrent with this decrease was an 
increase in pH. Subsequently, the 




250 300 



Figure 6. — Drainage A column effluent 

quality as a function of cumulative 
effluent volume. 



effluent CEU levels varied between 0.6 and 
0.7 mg/L and pH stabilized at about 7. 
Some problems were encountered in main- 
taining flow rates, and were attributed to 
compaction of the peat-sand bed. 



DISCUSSION 

In a wetland setting, peat can provide 
a reasonably rapid removal of trace metals 
from mine drainage for an extended period. 
The removal kinetics are affected by 
drainage composition. To reach equili- 
brium, the peat-metal reaction required 



296 



less than one hour with the acidic drain- 
age and roughly 24 to 48 hours for the 
neutral drainages. The slower kinetics 
observed for Leachates A and B may have 
been influenced by several factors. First, 
the decreased peat loading and the lower 
metal concentrations would tend to slow 
apparent reaction rates, assuming typical 
relationships between rates and reactant 
concentrations. Secondly, dissolved 
organics which have leached from the peat 
may form soluble organometallic complexes 
and thereby effectively compete with 
functional groups on the peat for low 
concentrations of trace metals. Fulvic 
acid, for example, are present in peat and 
are, by definition, soluble in both acidic 
and basic solutions. Gamble et al. (1970) 
indicated that these ill-defined compounds 
can chelate copper. Thirdly, the in- 
fluence of other cations, such as calcium, 
competing for binding sites on the peat 
would tend to be greater at low trace 
metal concentrations. Trace metals have a 
greater tendency than calcium to complex 
with organics. However, calcium might 
compete with some effectiveness since its 
concentration exceeded that of the trace 
metals by as much as four orders of magni- 
tude. Even considering the slower rate, a 
retention time of 24 to 48 hours does not 
seem unreasonable for a wetland system 
treating mine drainage. A travel time of 
50 days was reported for the flow of mine 
drainage through a cedar swamp in north- 
eastern Minnesota (Eger et al. 1980). 

The degree to which trace metal 
concentrations were reduced simulates the 
treatment efficiency in a field setting. 
The maximum metal removal from Drainages A 
and B, or treatment efficiency, typically 
was 60 to 97 percent of the aqueous metal. 
The fraction of nickel removed was great- 
est, while that of cobalt and zinc was the 
least . With Drainage C the copper removal 
was greatest (80 pet), while removal of 
Ni, Co, and Zn ranged from 35 to 40 
percent . 

Analyses of solution composition 
revealed the extent to which aqueous 
concentrations were reduced, but did not 
quantify the potential lifetime of peat in 
a field setting. One objective of the 
isotherm experiments was to generate data 
required for application of the Langmuir 
isotherm equation, and thereby determine 
removal capacities. With the exception of 
nickel, Langmuir isotherms did not des- 
cribe the metal removal adequately. 
Similarly, Ryss and Hoffmann (1979) 
reported nonlinear Langmuir results for 
peat with Cu , Ni, Cd, and Zn . This was 
not totally surprising since this applica- 
tion of Langmurian theory assumes removal 
is the result of adsorption by a homo- 
geneous surface from a simple solution. 
Both the peat and the drainages were 
complex in composition, and mechanisms 
other than adsorption were probably in- 
volved in removal reactions. At low 
concentrations, the complexation of trace 
metals by organics leached from the peat 



could have a significant influence on 
adsorption. The analytical error at the 
low concentrations, particularly for 
cobalt and zinc, is also greater. Since 
Langmurian theory did not describe the 
data, the theoretical removal capacities 
could not be calculated. Therefore, metal 
sequestration could be quantified only in 
terms of observed removal. 

Wieder (1984) used the Langmuir 
equation to describe metal sorption onto 
peat, but his experimental approach dif- 
fered in several aspects. First, he used 
a flow through system rather than batch 
reactors. Second, his solutions contained 
ferrous iron concentrations as high as 200 
mg/L, as opposed to the relatively low 
trace metal concentrations typical of this 
study. Third, his solutions were prepared 
from ferrous ammonium sulfate, as opposed 
to more complex drainages used in these 
batch studies. Fourth, as opposed to the 
linear transformations used to calculate 
Langmuir constants in this study, he used a 
nonlinear technique, which appears to be 
superior (Rubin and Mercer 1981). More 
recently Wieder (1988) found that the 
nonlinear estimation method adequately 
determines the Langmuir adsorption capacity 
but yields questionable values for the 
energy of interaction for peat adsorption 
of metals from more complex solutions. 

The mass of trace metals removed was 
influenced by aqueous metal concentrations 
and increased with increasing pH. The 
greatest mass removal was observed for 
nickel, which was the predominant trace 
metal in solution (table 3). Trace metal 
concentrations in Drainage B were higher 
than those in Drainage A. Similarly, trace 
metal removal from Drainage B, was greater 
than that from Drainage A. Although metal 
concentrations in Drainage C were higher 
than those in the other two drainages , the 
metal sequestration was not always higher. 
This was due in part to the pH of Drainage 
C, which was 4.2 units lower than that of 
the other drainages. The acidic conditions 
reduced the removal efficiency of the peat, 
as was verified in batch tests with Drain- 
ages A and B. This suggests that metal 
removal by peat in a wetland treatment 
system will be enhanced if the pH of acidic 
drainages is elevated prior to input to the 
wetland. Elevation of pH could also be 
accomplished by addition of alkaline solids 
to the peatland, but this could lead to 
chemical precipitation and possibly fouling 
of the peat. Ground water input to the 
peatland can provide some natural pH 
elevation . 

Reduced removal at low pH was also 
observed in the column experiments. With 
Drainage A, tailings were required to 
neutralize the acid present on the peat. 
The alkalinity of Drainage B was three 
times that of Drainage A, and effectively 
neutralized the acidity of the peat. The 
initial effluent from both types of columns 
had relatively low reductions in trace 
concentrations and depressed pH . The pH 



297 



Table 3. — Metal binding by peat, values in mg/kg. 



Metal Batch Experiments 

Drainage A Drainage B Drainage C 



Column 
Drainage B 



Cedar, 
swamp' 



Cu 
Ni 
Co 
Zn 
Fe 
Mn 



33U 

800 

2 

200 

NA 

NA 



730 

26,000 

110 

900 

NA 

NA 



2,500 

4,000 

300 

200 

NA 

NA 



29 


3,600 


18,000 


6,400 


580 


3,600 


890 


270 


NA 


26,000 


NA 


15,000 



NA : not analyzed. 

^Maximum values observed based on reductions in aqueous concentrations. 
Average values for three columns; columns approached, but did not reach 
breakthrough. Drainage A columns excluded since it was difficult to 

-discriminate between metal bound by peat and that bound by the tailings. 
Maximum observed total extractable metal concentration. 



Table 4. — Geochemical enrichment factors in the laboratory and field, 



Metal Batch Experiments 

Drainage A Drainage B 



9,100 

3,000 

330 

4,100 

NA 
NA 



Cu 


7,200 


Ni 


2,600 


Co 


69 


Zn 


2,000 


Fe 


NA 


Mn 


NA 





Column 


Cedar 


inage C 


Drainage B 


swamp 


23 


1,400 


5,800 


9.3 


3,300 


360 


12 


3,400 


3,000 


12 


5,200 


270 


NA 


NA 


19,000 


NA 


NA 


2,300 



NA : not analyzed. 



depression was also observed in the batch 
experiments and was not surprising since 
the peat pH was 4.5. In addition to 
hydrogen ion equilibration between peat and 
water, hydrogen ions present in certain 
functional groups, for example carboxyl 
groups, would be released in ion exchange 
reactions involving cations in solution. 

Geochemical enrichment factors (GEF, 
the ratio of the metal concentration in 
the peat to metal concentration in solu- 
tion) normalized for differences in 
aqueous metal concentrations provided the 
closest simulation of field values. The 
GEF values calculated for the neutral 
drainages typically ranged between 1,400 
and 7,200. The GEF values for copper and 
cobalt in a white cedar swamp receiving 
neutral mine drainage were in a similar 
range ( table 4 ) . 

The nickel and cobalt GEF values 
observed in the field were lower than this 
range. This may be partly due to the fact 
that the values listed in table 4 tend to 
underestimate the actual GEF values. The 
tabulated values were calculated using the 
concentrations at the stockpile, which 



were higher than those at the site where 
the metal content of the peat was analyzed. 
Secondly, equilibrium with respect to these 
metals may not have been achieved in the 
swamp. Finally, the GEF value for zinc at 
the low aqueous concentrations observed in 
the field may tend to be aberrant. At 
these low concentrations there may be 
effective competition from aqueous phase 
complexing organics leached from the peat. 
Extremely low GEF values were also observed 
for cobalt in the batch tests with Drain- 
ages A and B. 

Peat has been reported to sequester 
some trace metals preferentially over 
others (Bunzl et al. 1976, Ryss and 
Hoffmann 1979, Schmitzer and Khan 1972), 
and results of this study suggest that 
copper is preferred over nickel, cobalt and 
zinc. With Drainage C copper was preferen- 
tially removed, and copper concentrations 
in Drainages A and B were reduced to low 
equilibrium values at low peat loadings. 
The highest calculated GEF values were for 
copper, except in column experiments with 
Drainage B. In this experiment aqueous 
copper concentrations were very low, and 
the previously mentioned competition from 



298 



aqueous phase organics may have been 
influential. Field data from other north- 
eastern Minnesota sites indicate a pre- 
ferential sequestration of copper by 
organics. Analysis of organic stream 
sediments contacting mine drainage yielded 
GEF values for copper, nickel, cobalt, and 
zinc of 25,000; 3,500; 8,300; and 3,800, 
respectively (Eger and Lapakko 1980). 

The results of batch and column 
experiments indicate that solution compo- 
sitional factors such as pH, metal concen- 
tration, and presence of competing metals 
influence trace metal removal. The com- 
position and hydraulic conductivity of 
peat and wetland hydrology also influence 
trace metal removal. Consequently, the 
efficiency and capacity of peat to remove 
trace metals from specific drainages can 
be accurately assessed only on a site 
specific basis. Nonetheless, approxima- 
tions can be used to gain some insight on 
the amount of metal removal a peatland can 
provide . 

The batch and column experiments 
indicated that one kilogram of peat (dry 
weight) is capable of sequestering roughly 
20 g of nickel from Drainage B, field data 
suggest a value of slightly more than 6 g. 
For this example a reasonable and conven- 
ient nickel removal capacity of 10 g/kg 
will be used. For field calculations re- 
moval capacity might be more appropriately 
expressed as 10 kilograms nickel per 
metric ton of peat. Assuming a peat depth 
of 1 m and a dry density of 0.1 tons/m , 
this implies a removal capacity of 10 tons 
of nickel per hectare of peatland. 

In a natural peatland problems are 
presented by the low hydraulic conduc- 
tivity of peat (Rycroft et al. 1975). 
Flow does not occur uniformly throughout 
the peat , but instead tends to occur over 
the surface and in the upper peat layer 
(Eger, et al. 1980). Consequently, it is 
advantageous to modify the peat layer in 
wetlands receiving mine drainages. 
Furrowing or disking the peat (perpen- 
dicular to the direction of flow) would 
allow the drainage to contact a greater 
surface area of peat and, thereby, improve 
metal removal. Flow through the lower 
layers of peat may also be enhanced by 
ditching perpendicular to the natural flow 
gradient. The effectiveness of this 
approach may, however, be limited by the 
slight gradients which are typical of 
natural peatlands. If peat were removed 
from the natural setting, the hydraulic 
conductivity could be increased by mixing 
the peat with inorganic solids, such as 
the sand or tailings used in the column 
experiments. Use of an alkaline material 
such as limestone, with this or other 
approaches, would further enhance trace 
metal removal by elevating pH. 

Based on the desorption experiments 
with distilled water, the metals removed 
by the peat would remain tightly bound if 
the peat were contacted by precipitation. 



From 1983 through 1986 the volume weighted 
average precipitation pH at four sites in 
northeastern Minnesota ranged from 4.62 to 
5.21, with a median of 4.82 (Twaroski, 
1988). The degree of release would in- 
crease with decreasing pH. Indeed, it is 
possible that metals could be recovered 
from the peat when its capacity was 
reached. Metals could be released from the 
sorbed state by acidic solutions or by 
solutions containing strong complexing 
agents, as inferred by the elevated metal 
release to bog water. Exposing the metal 
loaded peat to high ionic strength solu- 
tions would also release bound metals. The 
batch experiments indicated that the acidic 
solution and bog water were superior for 
desorbing trace metals, but more efficient 
salt solutions than that used in these 
tests most likely exist. Another option 
would involve burning the spent peat, with 
possible attendant energy recovery, and 
recovering metals from the ash. The 
surface peat could be removed as its 
capacity was exhausted, thereby exposing 
the less taxed lower layers of peat. 



SUMMARY 

Laboratory studies indicate that peat 
is capable of reducing trace metal concen- 
trations in mine drainage by more than 90 
percent, and the reaction kinetics are 
adequate for peat to be a major trace metal 
sink in wetlands. The trace metal removal 
capacity is on the order of 20 g/kg, is 
dependent on trace metal concentrations, 
and decreases as pH decreases. At low 
concentrations, removal may be limited due 
to trace metal complexation by dissolved 
organics which leach from the peat. The 
batch and column data provided a reasonable 
simulation of trace metal removal in the 
field when the influence of solution 
composition was considered. If wetland 
systems are engineered to permit the 
contact of mine drainage with lower layers 
of peat, tens of tons of trace metals can 
be removed per hectare of peatland. 



LITERATURE CITED 

Bunzl, K. , W. Schmidt, and B. Sansoni. 
1976. Kinetics of Ion Exchange in 
Soil Organic Matter. IV. Adsorption 
and Desorption of Pb(II), Cu(II), 
Zn(2), and Ca(II) by Peat. J.- Soil 
Sci. , V. 27, pp. 32-41. 

Eger, A. P., K. Lapakko, and P. Otterson. 
1980. Trace Metal Uptake by Peat: 
Interaction of a White Cedar Bog and 
Mining Stockpile Leachate. Proc. 6th 
International Peat Cong. , 
pp. 542-547. 



299 



Eger, A. P. and K. Lapakko. 1980. 

Transport of Chemical Constituents 
Present in Mining Runoff through a 
Creek System. MN Dep . Nat. Resour. 
Minerals Div. St. Paul, MN, 47 pp. 
plus appendices (appendix 14). 

Fraser, D. C. 1961. Organic Sequestration 
of Copper. Econ. Geol., V. 56, 
pp. 1063-1078. 

Gamble, D. S., M. Schnitzer, and I. 

Hoffman. 1970. Cu(II) Fulvic Acid 
Chelation Equilibrium in 0.1 M KC1 at 
25° C. Can. J. Chem. , V. 48, 
pp. 3197-3204. 



Lapakko, K. A., J. D. Strudell, and 

A. P. Eger. 1986. Low-Cost Removal 
of Trace Metals from Copper-Nickel 
Mine Stockpile Drainage. Volume II, 
Trace Metal Sequestration by Peat, 
Other Organics, Tailings and Soils: 
A Literature Review (Contract 
J0205047). BuMines NTIS # PB 87- 
186144, 45 pp. 

Lind, D., K. Alto, and S. Chatterton. 

1978. Regional Copper-Nickel Study, 
Aquatic toxicity study. MN Envir. 
Quality Board, St. Paul, MN , 54 pp. 



linnesota Pollution Control Agency, 
Personal communication with 
Dave Maschewitz of the MPCA. 



1988. 



Ong, H. L., and V, E. Swanson. 1966. 
Adsorption of Copper By Peat, 
Lignite, and Bituminous Coal. Econ. 
Geol., V. 61., pp. 1214-1231. 

Oreshko , V. F. , A. I. Berdnikov, and 
V. V. Prezhbylskii. 1962. The 
Sorption of Radioactive Cobalt on 
Peat. Radiokhimiya 4, pp. 499-502. 

Premi, P. R. 1970. Effect of Different 
Levels of Copper and Zinc Added as 
Adsorbed Cations on Purified Peat and 
Sewage on Nitrogen Mineralization 
During Incubation of Soil. 
Agrochem. , V. 14, 372 pp. 

Rubin, A. J. and D. L. Mercer. 1981. 
Adsorption of Free and Complexed 
Metals from Solution by Activated 
Carbon. Tn M. A. Anderson and 
A. J. Rubin (eds.). Adsorption of 
Inorganics at Solid-Liquid Inter- 
faces. Ann Arbor Science Publishers, 
Inc. Ann Arbor, MI, pp. 295-324. 



Rycroft, D. W. , D. J. A. Williams, and 

H. A. P. Ingram. 1975. The Trans- 
mission of Water Through Peat. J. 
Ecol. 63, pp. 535-556. 

Ryss, K. A., and M. R. Hoffmann. 1979. 

Removal of Trace Metals from Aqueous 
Systems by Adsorption on Bog 
Material. Rep. to MN Environ. 
Quality Council, Regional Copper- 
Nickel Study. State Planning 
Agency, St. Paul, MN, 21 pp. 

Schnitzer, M. , and S. U. Khan. 1972. 

Humic Substances in the Environment. 
Dekker, 327 pp. 

Smith, E. F., P. MacCarthy, T. C. Yu and 
H. B Mark. 1977. Sulfuric Acid 
Treatment of Peat for Cation Ex- 
change. J. Water Poll. Control Fed., 
v. 49, pp, 633-638. 

Szabo, I. 1958. Adsorption of Cations on 
Humus Preparations. Comm. Third 
Math-Phys. Class Hung. Acad. Sci., 
V. 8, pp. 393-402. 

Twaroski, C. 1988. Personal communica- 
tion with Cliff Twaroski of the 
Minnesota Pollution Control Agency 
Acid Precipitation Program. 
St. Paul, MN. 



U.S. 



Environmental Protection Agency. 
1980. Ambient Water Quality Criterij 
for Zinc - 1980. EPA 44/5-80-079, 
67 pp. 

1984. Ambient Water Quality Cri- 
teria for Copper - 1984. EPA 440/ 
5-84-031, 142 pp. 

1986. Ambient Water Quality Cri- 
teria for Nickel - 1986. EPA 440/ 
5-86-004, 93 pp. 



Wieder, K. 1984. Personal communication 
with Kelman Wieder of the Villanova 
University Biology Department. 
Villanova, PA. 



1988. Personal communication with 
Kelman Wieder of the Villanova 
University Biology Department. 
Villanova, PA. 



300 



NICKEL AND COPPER REMOVAL FROM MINE 
DRAINAGE BY A NATURAL WETLAND 



Paul Eger and Kim Lapakko' 



Abstract. — Recently a considerable amount of 
attention has been focused on the use of wetlands, 
both natural and constructed, for the treatment of 
acid coal mine drainage. Wetlands also may have a 
significant capacity for removing trace metals from 
other types of mine drainage. Minnesota has many 
small and large peatlands near existing and poten- 
tial mining developments, and the use of peatlands 
for the control of drainage quality is an attractive 
alternative to chemical treatment. A study has been 
conducted on a white cedar peatland receiving 
stockpile drainage which had an average concentra- 
tion of 17.9 mg/L nickel and 0.62 mg/L copper. 
Surface and ground water, vegetation (white cedar 
( Thuja occidentalis ) , alder (Alnus rugosa ) , sedge 
(Carex spp . ) ) and peat were all analyzed for trace 
metal content. Based on mass balance calculations, 
and water quality data, essentially all of the 
copper and 80% of the nickel were removed by the 
peatland. Removal by peat accounted for greater 
than 90% of the overall metal reduction. Maximum 
observed concentrations in the peat were 6400 mg/kg 
nickel and 3600 mg/kg copper; while the maximum 
values for the vegetation were 239 mg/kg nickel and 
10.1 mg/kg copper. 



Paper presented at the 1988 Mine Drainage and Surface Mine Reclamation Conference 
sponsored by the U.S. Department of the Interior (Bureau of Mines and Office of Surface 
Mining Reclamation and Enforcement), April 17-22, 1988, Pittsburgh, PA. 

"'Paul Eger is Principal Engineer and Kim Lapakko is Senior Engineer, Minnesota Department 
of Natural Resources, Division of Minerals, St. Paul, MN . 



301 



INTRODUCTION 

Wetlands, both natural and constructed, 
have been used successfully in the treat- 
ment of acid mine drainage (Kleinman, 
1985; Wieder et al., 1982). Wetlands and 
the peat they contain have a significant 
trace metal adsorption capacity and offer 
a potential low maintenance treatment 
technique for metal laden mine drainage. 
In laboratory studies, peat has been used 
successfully to remove trace metals from 
various types of stockpile drainage 
(Lapakko and Eger, 1981, Lapakko et al., 
1986). 

Minnesota has 6 million acres of peatland 
most of which are located in the northern 
portion of the state. This area also has 
the highest mineral potential and explora- 
tion for gold and base metals is presently 
occurring on over 250,000 acres. Wetland 
treatment systems for mine drainage could 
be applicable to mining developments in 
Minnesota . ' 

The objective of this study was to deter- 
mine both the short-term and long-term 
effectiveness of a peatland in removing 
trace metals from stockpile drainage. 
This paper describes the initial phase of 
the study; the evaluation of the short- 
term effectiveness. 




t;::M: : 3 Stockpiles 
1*7*71 Peatland 



SITE DESCRIPTION 



Figure 1. — Dunka Mine, stockpile and 
sampling locations. 



The initial phase of the study was 
conducted from July 1976 to August 1977 at 
the LTV Steel Mining Company's Dunka Mine 
in northeastern Minnesota. The Dunka Mine 
is a large open pit taconite operation, 
covering approximately 160 ha. The pit is 
4 km long, .4 km wide and has a maximum 
depth of 110 meters. At this location, 
the Duluth complex, an igneous intrusion 
overlies the taconite ore and must be 
removed and stockpiled. The material has 
been separated based on copper content and 
has been stockpiled along the east side of 
the open pit. Drainage from all stock- 
piles and mine dewatering discharges (011, 
012) flow to Unnamed Creek (fig. 1). 

Drainage from a stockpile containing 
Duluth complex material with an average 
grade of 0.30% copper and 0.09% nickel 
flowed through a white cedar ( Thuja 
occidentalis ) peatland into Unnamed Creek 
(fig. 2). The stockpile covers 0.12 km ; 
flow rates varied from zero during the 
winter to a maximum of 16 L/sec. Mean 
trace metal concentrations were 17.9 mg/L 
nickel, and 0.62 mg/L copper; the mean pH 
of the drainage was 7.2 (table 1). 



. » OT-9 
- ■ • « CEDAR 

ySEEPJ "vXAViV DT-10/ 

' ■"■>—'"' "dt-11 . »'» 

V 3-1 "^ ' » * 

T-1^ 03-2 T-12Q\ k \\\\\ 

TV2D 03-3 aT _, 3 N ^V'*« 




AEM-6 



MONITORING SITES 
O Well 
D Terrestrial 
A Stream 



60 m 



The peatland covers 0.04 km , the depth of 
the peat ranges from 1.5 to 1.8 m, and the 
area is covered by 3-5 cm of standing 
water. The peat is generally well- 
decomposed, with decomposition increasing 
with depth, and is underlain by a layer of 
silty blue clay. Another white cedar 



Figure 2. — White cedar peatland. 



302 



Table 1.— Stockpile Drainage Water Quality, July 1976 through August 1977. 



pH 

Alkalinity (mg/1 as CaOCL) 

Specific conductance (us/cm) 

Calcium (mg/1) 

Magnesium (mg/1) 

Sulfate (mg/1) 

Copper (mg/1) 

Nickel (mg/1) 

Cobalt (mg/1) 

Zinc (mg/1) 

Iron (mg/1) 



Mean 



Range 



7.2 


6.68 - 7.79 


24 


113 


47 - 206 


24 


2540 


890 - 3550 


25 


285 


93 - 388 


17 


225 


52 - 288 


12 


1300 


370 - 2600 


17 


.62 


.04 - 1.7 


24 


17.9 


.4 - 39.8 


24 


1.16 


.4 - 2.4 


8 


.38 


.07 - .65 


12 


1.44 


.4 - 5.4 


24 



peatland in the same watershed but remote 
from trace metal sources, was selected as 
a control. 



METHODS 

Sampling stations were established at 6 
sites along the stream, at each of the 
major stockpile flows (Em-8, Seep 1, Seep 
3), at the two dewatering discharges (011, 
012), and at 26 sites in the peatland 
(figs. 1, 2). Continuous flow data was 
collected at three sites along the stream 
(Em-1, Em-3, Em-5) at the largest volume of 
stockpile seepage (Em-8), and pumping 
records were available for the mine 
dewatering discharges. Flow measurements 
at all other stream and stockpile sites 
were collected every two weeks. 

Water quality samples were collected at 
each site twice monthly during open water 
(typically April 1 - November 15) and 
approximately monthly during the winter. 
There was no drainage from any of the 
stockpiles during the winter. 

In addition to the twice monthly samples 
special sampling programs were conducted to 
better quantify the source of metal input 
and the amount of metal transport in 
various parts of the watershed. Automatic 
samplers were used to collect composite 
samples along the stream, low flow samples 
were collected under the ice, and dye 
studies with Rhodamine WT were conducted to 
determine the travel times between the 
different sampling stations on the streams. 
Water quality samples were collected 
sequentially (from upstream to downstream) 
at the specific time intervals measured in 
the dye study so that the same parcel of 
water was sampled as it moved downstream. 

In the peatland, at 8 of the sites (3-1 to 
3-8) shallow wells and piezometers were 
installed. The shallow wells consisted of 
a perforated 10 cm diameter, 46 cm long PVC 
pipe, taped at the top to minimize surface 
water infiltration. The piezometers (1.5 m 
depth) were constructed of PVC, with a 
flange and a bentonite seal to prevent 



leakage. Specific conductance was sampled 
in the surface water at all 26 sites, and 
in the piezometers and wells. Specific 
conductivity surveys of the surface water 
were conducted monthly during the summer of 
1977 to establish the path of stockpile 
seepage through the peatland. All 26 sites 
and additional sites perpendicular to the 
well line were sampled. In June and 
August, at sites 3-1 to 3-8, the water at 
the surface and in the wells and piezo- 
meters was filtered, and analyzed for trace 
metals by atomic absorption. 

At the odd numbered T sites and at the 
control peatland within the same watershed, 
releves (Mueller-Dombois and Ellenberg, 
1974) were used to describe the plant 
community; and visual estimates of plant 
damage were made. Samples of leaf tissue 
were collected at each T site for each of 
three species; white cedar ( Thuja 
occidentalis ) , alder ( Alnus rugosa ) , and 
Carex spp. . These samples were wet ashed 
using a HN0 3 /HC10 3 digestion and analyzed 
for trace metal content using inductively 
coupled plasmaspectroscopy (ICP). Diges- 
tions and analyses were performed by 
Barringer Laboratories in Toronto, Canada. 

At each site a composite of three peat 
samples was collected from the top 20 cm 
and analyzed for trace metals by total acid 
extraction (HF, HC1, HN0„) (Meineke and 
Klaysmat, 1976). At fouf stations (3-1, 
3-2, 3-5, 3-8) and the control site, trace 
metal concentrations were determined at 
approximately 25 cm depth intervals for the 
entire depth of the peat. 



RESULTS 

Stockpile drainage typically began in late 
March and ended around the middle of 
November, while mine dewatering and ground- 
water inputs maintained flow in the stream 
for the entire year. The mean nickel 
concentration at the mouth of the stream 
(site Em-1) was 0.1 mg/L and ranged from 
0.03 to 0.22 mg/L. Concentrations through- 
out the stream exceeded the natural back- 
ground levels (.001-. 005 mg/L) for streams 



303 



Table 2. — Water Quality Summary at White Cedar Peatland, August 1977. 



parameter 



surface samples 
mean range 



shallow wells 
mean range 



piezometers 
mean range 



specific conductance 


3100 


2250 - 


3500 


2600 


2100 - 


3000 


270 


213 - 


325 


(us/cm) 




















copper (mg/1) 


.13 


.002 - 


.46 


.001 


.001 - 


.002 


.001 


.001 - 


.002 


nickel (mg/1) 


20 


8.4 - 


26.0 


.6 


.18 - 


.93 


.03 


.01 - 


.06 



in the area, even during the winter when 
stockpile input ceased. Copper concentra- 
tions ranged from .001 to .006 mg/L and 
were similar to the values measured in 
unimpacted streams (.001-. 005 mg/L) 
(Thingvold et al., 1979). 

Water quality in the peatland varied with 
distance from the drainage input , depth in 
the peatland, and time. The specific 
conductance values in the surface water 
were used to define the path of stockpile 
drainage through the peatland. The 
boundary on the west side was particularly 
dramatic; specific conductance decreased 
from over 2000 to 400 within several 
meters. Specific conductance decreased as 
distance from the seep and depth in the 
peatland increased. Conductivity values 
decreased from 3000-3500 us/cm at the seep 
to 600-2000 us/cra near the stream with the 
majority of the reduction occurring over 
the final 200 meters (fig. 3). Specific 
conductance values were highest in the 
surface water and were an order of magni- 
tude greater than those measured in the 
piezometers (table 2). 

The reduction of copper and nickel concen- 
trations with increasing distance from the 
seep in surface samples was greater than 
that of specific conductance (fig. 3). In 
both June and August the copper concentra- 
tion was reduced by about two orders of 
magnitude. The greatest reduction 
occurred within the initial 100 meters of 
flow, with concentration remaining 
relatively constant, essentially at or 
slightly above background levels ( 4 .005 
mg/1) over the final 200 meters. In June, 
nickel was reduced by about two thirds 
within the first 150 meters, but in August 
little reduction occurred in this portion 
of the peatland. In August, nickel did 
decrease by about two thirds over the 
final 200 meters of the peatland but the 
concentration near the stream (8.4 mg/L) 
was still about three orders of magnitude 
greater than background values. Surface 
concentrations of both metals were two to 
three orders of magnitude greater than the 
values from the wells and the piezometers 
(table 2) . 



Vegetation 

The predominant vegetation in both the 
impacted and control peatlands was white 



> *sfc 




SPfCIPIC CONOUCTAMCt 




DISTANCE FROM SEEP, R 



Figure 3. — Copper, nickel, and specific 

conductance in surface water in white 
cedar peatland, June and August, 1977. 

cedar, alder and Carex. Damage levels were 
determined by the presence of chlorosis in 
the leaves, and gave only a qualitative 
indication of plant stress. High levels of 
damage to vegetation were observed at two 
of the ten T sites (T T ), but no 
definitive pattern relating damage to 
distance from the seep was observed. 

Leaf tissue nickel concentrations were 
about one to two orders of magnitude higher 
in the impacted area than the values 
measured in the control peatland, where 
nickel was not detectable ( <1 mg/kg). The 
highest nickel concentrations were 
generally found within the zone of high 
conductance water and the lowest values 
were measured outside this zone in the 
northeast section of the peatland at sites 
T 9 _T H (fig. 4). Elevated nickel values 
were, however, found at sites T -T 1Q which 
were outside the zone of high specific 
conductivity. Copper concentrations were 
about an order of magnitude lower than the 
nickel values even at sites near the seep, 
and there was no relationship between the 



304 



KIT: WKitt C«fl«r ( T h u } • oc c id*»< ftllal 

C«r«a <••■.) 



I C 

I cut u 

> •• 

\ 31 

. otii 



OH 



O'oo "» v 



x Zona of 

x high conductance 



\/ 



A\ \ 




Figure 4. — Nickel concentration in 

vegetation, mg/kg (August 1977) 



copper values and location in the peat- 
land. Concentrations in the leaves of 
cedar and alder were slightly higher at 
the control than the values measured at 
the impacted area (table 3). 

Peat 

Nickel and copper concentrations in the 
peat decreased as distance from the seep 
and depth in the peat column increased. 
The maximum copper concentration of 3600 
mg/kg was found in the top 20 cm sample 
near the seep (site 3-1). The concentra- 
tion decreased with depth and the maximum 
value was about 6 times the value at the 
base of the peat (fig. 5). Samples 
collected at sites further from the seep 
had a more uniform concentration versus 
depth profile and concentrations near the 
creek were only 200-400 mg/kg. Surface 
nickel concentrations ranged from 6400 



mg/kg near the seep to 1116 mg/kg near the 
creek and decreased with depth at all sites 
in the impacted area (fig. 6). Isopleths 
of the concentrations in the top 20 cm and 
the specific conductance in the surface 
water were used to define the area of 
drainage impact (fig. 7). 

Generally the zone of high specific 
conductance corresponded to the area with 
the highest nickel concentrations in the 
peat. However, concentrations increased 
along the line from T-g-T and the nickel 
values at T ig was comparable to the values 
within the zone of influence. Concentra- 
tions of the trace metals at the control 
peatland (25 mg/kg average copper, 65 mg/kg 
average nickel) and across the stream (site 
3-8) were one to two orders of magnitude 
lower than the concentrations near the 
stockpile and were relatively constant with 
depth at both sites. 

Mass Balance 

To quantify the overall removal of copper 
and nickel occurring in the peatland a mass 
balance was calculated for the period of 
study. An overall watershed balance and a 
balance for the peatland alone were 
computed. Flow and concentration measure- 
ments were combined to compute overall mass 
inputs and outputs. Individual time 
periods were analyzed and error estimates 
for the nickel mass were made for Seep 3, 
Em-8 and Em-1. The overall mass values 
were estimated to be within + 15% (Eger and 
Lapakko, 1980). 

For the watershed, the major source of 
nickel and copper was the stockpile 
drainage, with seep 3 contributing about 
87% of the total nickel input and 78% of 
the total copper input (table 4). The 
total input from stockpiles comprised 99.2% 
and 82% of the nickel and copper inputs, 
respectively. The total nickel input was 
2087 kg and total output, as measured at 
the mouth of the stream (Em-1), was 340 kg. 
The overall nickel removal was 84%. Total 
copper input was 112 kg and the output was 
9 kg, resulting in an overall copper 
removal of 92%. 

The only significant input of nickel and 



Table 3. — Trace Metal Concentrations in Vegetation. 



Plant 
Species 



Peatland Receiving Stockpile Drainage 
(sites Tl - T19) 



Copper (mg/kg) 
mean range 



Nickel (mg/kg) 
mean range 



Control Peatland 

(mean, composite 

sample) 

Copper Nickel 
(mg/kg) (mg/kg) 



"Ttegional Values 



Copper 
(mg/kg) 



Nickel 
(mg/kg) 



No. of 
Sites 



White Cedar ( Thuja occidentalis ) 2.6 2.1 - 3.8 37 8-62 4.1 1.0 No Data No Data 1 
Alder ( Alnus rugosa ) 7.8 4.8-10.1 82 10-239 12.1 1.0 3.9-12.1 1.0 3 

Carex spp . 6.1 3.3-9.2 76 7-233 2.6 1.0 0.9-8.6 1.0-8.0 10 



TJinni 



esota Regional Copper-Nickel Study 



305 



I 

EDGE 
OF I 
WET- I 
LAND 



NO DATA 



27 SO 
EMS 3-1 



70 
EMS 3-2 



230 
EMS 3-5 



307 363 

EMS 3-8 



DISTANCE FROM SEEP (MEASURED ALONG LINE OF SAMPLING STATION), meters (NOT TO SCALE) 



Figure 5. — Copper concentration (mg/kg) vs 
depth and distance from seep. 



a 






6400 j 




SOOO 


1116 




56 


0.2S 












0.50 

0.7S 






2200 


863 


422 




60 






1170 


777 


366 




66 






















760 


366 


228 


III 


80 


1.00 






, 






K 
O 










662 


424 


326 




164 


1.28 






















644 


412 


342 




80 


1.60 


- 




















460 


367 


292 




NO DATA 
















1.76 

M 


EDGE 

.Vt- 

LAND 












■> 


1 _| 













27 50 
EMS 8-1 



70 
■MS 3-2 



930 
EMS 3-8 



307 363 

EM* 8-8 



DISTANCE FROM SEEP (MEASURED ALONG LINE OF SAMPLING STATION), meters (NOT TO SCALE) 



Figure 6. — Nickel concentration (mg/kg) vs 
depth and distance from seep. 



306 



Table 4. — Overall Mass Balance for Watershed. 





Volume^ 




Mass (kg) 






(L 


X 1<T) 


Nickel 


Copper 


Sulfate 


Inputs 












Stockpile Seepage 












Seep 3 




90 


1810 


72 


103,000 


En8 




184 


150 


4 


128,000 


Seep 1 




29 


110 


16 


72,000 


Mine Dewatering Pumps 












Oil 




2530 


14 


16 


205,000 


012 
ifatural Runoff 




255 


0.8 


2 


78,000 




426 


2 


2 


13,000 


Total Input 




3514 


2087 


112 


599,000 


Outflow 












Qn-1 




3514 


340 


9 


563,000 



•W 



erall Removal 



84 



92 



input - output X 100% 
input 

"Computed by difference. 
Volume at Fm-1 minus the sum of all input volumes. 



Table 5. — Nickel Mass Input from Stockpile 
Drainage and Release from the Peatland 
to Stream. 





Nickel 


mass loading 


Time Period 


Seep 
(mg/sec) 


Input to 

Stream from 

Peatland 

(mg/sec) 


7-15 to 10-25-76 

12-07 to 12-10-76 

5-05 to 5-13-77 

8-16 to 8-17-77 


45.4 


1.1 
111.0 


3.6 

1.2 

5.4 

28.0 



copper to the white cedar peatland was the 
stockpile drainage from Seep 3. Nickel 
input during the 13 month study period was 
1800 kg, and copper input was about 70 kg 
(table 4). The output from the peatland 
was much more difficult to quantify since 
there was no one single point where the 
drainage from the peatland entered the 
creek. The drainage occurred in a diffuse 
manner along the stream for about 250 
meters (determined by specific conductivity 
measurements). The output of nickel was 
determined through the overall balance for 
the watershed, by comparing upstream and 
downstream nickel loads in the stream, and 
through low flow sampling along the stream 
(Eger and Lapakko, 1980). During the study 
period, discharge rates of nickel from Seep 
3 ranged from during the winter to 111 
mg/sec, while contributions to the stream 
ranged from 1.2 to 28 mg/sec during the 



• ot 

h conductance 




Figure 7. — Nickel concentration (mg/kg) in 
top 20 cm of peat, August 1977. 



same period (table 5). The total estimated 
load from the peatland to the stream was 
about 300 kg, or about 80% of the nickel 
had been removed in the peatland. The 
copper output was not calculated since 
water quality data in the peatland and in 
the stream indicated that 1) copper concen- 
trations were reduced to background levels 
within 200 meters from the seep and 2) 
there was no measurable input of copper 
from the peatland to the stream. 



307 



Metal accumulat 
water, vegetati 
specific conduc 
the boundary of 
Since the upper 
about 90% water 
tions indicated 
water in the pe 
amount of metal 
water. Using t 
at the well sit 
between the sit 
the mass of met 
For nickel the 
of the metal be 
of the peatland 



ion was estimated in the 
on and peat, using the 
tance values to determine 
the impacted area, 
portion of the peat is 
and since field observa- 
about 3-5 cm of standing 
atland, a substantial 

could be stored in the 
he concentrations measured 
es and dividing the area 
es, estimates were made for 
al stored in the water, 
total was 42 kg, with 76% 
ing stored in the top 20 cm 



Biomass estimates were made for cedar, 
carex, and alder (Ohman and Grigal, 1985; 
Dyer, 1967) and the average concentrations 
for each species was used to estimate the 
mass of nickel in each species. The total 
mass was less than 6 kg with cedar 
containing about 70% of the total. 

The total mass of nickel contained in the 
top 20 cm of peat was calculated by using 
the area between the isopleths and 
applying the mean metal concentration to 
that volume of peat. Since a certain 
percentage of the metals were contained in 
the peat prior to the stockpile drainage a 
background correction was applied. The 
concentration at depths greater than 75 cm 
at the sites 3-2 and 3-5 was assumed to be 
representative of natural conditions. 
Estimates were made for the 20-75 cm depth 
based on the measured depth profiles and 
assuming that the isopleths at depth would 
be similar to those in the top 20 cm. 
Using these assumptions the nickel mass 
was estimated to be about 1150 kg in the 
top 20 cm and 350 kg between 20-75 cm. 
The mass of copper was calculated in a 
similar fashion and was estimated to be 77 
kg. 



than 90% of the metals were associated with 
the peat. Although the nickel and copper 
concentrations in the peat are on the order 
of several tenths of a percent , laboratory 
experiments with stockpile drainage have 
demonstrated that peat can accumulate up to 
2% nickel during a continuous column 
removal experiment (Lapakko et al . , 1986) 
and peat from the Tantramar Swamp in New 
Brunswick was found to contain as much as 
10 percent copper (MacDonald et al., 1976). 
Typical metal concentrations for peat in 
Minnesota are about two orders of magnitude 
less than the values measured in the 
impacted peatland (Grigal and Nord, 1983). 
Metal values above background have been 
found near zones of mineralization, but 
were generally an order of magnitude less 
than the concentration in the study area 
(Meineke et al . , 1977). 

Specific conductance and peat metal concen- 
trations indicate that the contact zone 
between the stockpile drainage and the 
peatland is confined to the upper portion 
of the peat, a zone referred to as the 
acrotelm, or the zone of active water 
movement (Romanov, 1968; Ingram, 1978). 
Since peat decomposition generally 
increases and hydraulic conductivity 
decreases with depth, only a small amount 
of the stockpile drainage will contact the 
lower levels of the peat and therefore 
treatment is restricted to the upper more 
permeable zone. 

Even though the nickel values in the 
vegetation were elevated the removal of 
nickel by plant up take accounted for less 
than 1% of total nickel removal. Increas- 
ing the biomass in a wetland might provide 
greater metal removal, but for the metals 
in this study the increased removal would 
not appear to be significant. 



CONCLUSIONS 



DISCUSSION 

Significant removal of nickel and copper 
has occurred in this watershed. The 
overall watershed balance indicated that 
about 1750 kg of nickel and 100 kg of 
copper were being removed. The largest 
source of metal input was Seep 3, but 
analyses of upstream and downstream loads 
in the stream revealed that only a portion 
of the nickel and none of the copper were 
transported through the peatland. Most of 
the metal removal in the watershed was 
occurring in the white cedar peatland. 



Nickel and copper have been successfully 
removed from stockpile drainage as it 
flowed through a natural white cedar 
wetland. Essentially all of the copper and 
about 80% of the nickel was removed during 
the study period. Peat uptake accounted 
for over 90% of the removal while vegeta- 
tion uptake provided less than 1% of the 
overall removal. Wetland treatment appears 
to offer a low maintenance alternative to 
seepage collection and treatment of metal 
contaminated drainage. 



Both the input - output cal 
accumulation estimates indi 
about 1500 kg of nickel and 
copper were removed and sto 
peatland during the period 
Based on the overall mass b 
water quality results, esse 
the copper and about 80% of 
were removed. Analysis of 
compartments demonstrated t 



culations and 
cated that 

about 70 kg of 
red in the 
of study, 
alance and 
ntially 100% of 

the nickel 
the individual 
hat greater 



LITERATURE CITED 

Dyer, R. F., 1967. Technical bulletin #28, 
Maine Agricultural Research Station. 

Eger, P., and K. Lapakko, 1980. Transport 
of chemical constituents present in 
mining runoff through a creek system, 
Minnesota Department of Natural 
Resources, St. Paul, MN. 



308 



Grigal, D. F. , and W. S. Nord, 1983. 
Inventory of heavy metals in 
Minnesota peatlands. Department of 
Soil Science, University of 
Minnesota, St. Paul, MN. 

Ingram, H. A. P., 1978. Soil layers in 
mires: function and terminology. 
Journal of Soil Science 29: 224-227. 

Kleinmann, R. P., 1985. Treatment of acid 
mine water by wetlands. Control of 
Acid Mine Drainage: Proceedings of a 
Technology Transfer Seminar. U.S. 
Bureau of Mines Information Circular 
9027. 



Thingvold, D. , P. Eger, M. Hewett, B. 
Honetschlager, K. Lapakko , and R. 
Mustalish. 1979. Minnesota regional 
copper-nickel study, Vol. 3, Ch. 4. 
Water Resources, Minnesota Environ- 
mental Quality Board, St. Paul, MN. 

Wieder, R. K. , G. E. Lang, and A. E. 

Whitehouse. 1982. Modification of 
acid mine drainage in a fresh water 
wetland. Acid Mine Drainage Research 
and Development, 3d WV Surface Mine 
Drainage Task Force Symp. WV Surface 
Mine Drainage Task Force, Charleston, 
WV, pp. 38-62. 



Lapakko, K. , and P. Eger. 1981. Trace 
metal removal from mining stockpile 
runoff using peat, wood chips, 
tailings, till, and zeolite. 
Symposium on Surface Mining 
Hydrology, Sedimentology and Reclama- 
tion, Lexington, KY. 

Lapakko, K. , P. Eger, and J. Strudell, 
1986. Low-cost removal of trace 
metals from copper-nickel mine stock- 
pile drainage, Volume 1, Laboratory 
and Field Investigation, U.S. Bureau 
of Mines Mining Research Contract 
Report . 

MacDonald, R. J., K. E. Hague, and J. E. 
Dutrizoc. 1976. Copper recovery 
from copper-bearing peat moss. 
Miner. Res. Program, Miner. Science 
Lab. Report No. MRP/MSC 76-275 (1R). 

Meineke, D., and A. Klaysmat, 1976. 
Preliminary report on nineteen 
digestion methods tested on various 
geochemical exploration sample 
medias, Minnesota Department of 
Natural Resources, Report 104, 
Hibbing, MN. 

Meineke, D. G. , M. K. Vadis, and A. L. 

Klaysmat. 1977. Pilot study on peat 
exploration geochemistry, Birch Lake 
area, Lake County, Minnesota, 
Minnesota Department of Natural 
Resources, Report 108-1, St. Paul, 
MN. 

Mueller-Dombois, D., and H. Ellenberg, 

1974. Aims and methods of vegetation 
ecology. 547 p. John Wiley and 
Sons, NY. 

Ohman, L. F., and D. F. Grigal. 1985. 
Biomass distribution of unmanaged 
upland forests in Minnesota. Forest 
Ecology and Management 13: 205-222. 

Romanov, V. V., 1968. Hydrophysics of 

Bogs , Israel Program for Scientific 
Transactions . 



309 



EFFECTS OF A SPHAGNUM PEAT ON THE QUALITY OF A SYNTHETIC ACIDIC MINE DRAINAGE 



Jonathan M. Dietz and Richard F. Unz 



att r 

the 

unde 

and 

cone 

rece 

moni 

were 

sulf 

f err 

flux 

acco 

tran 

cond 

com 

i.e. 

the 

cone 

of s 

i n t 

iron 

add i 

Deer 

asso 

impo 

wet 1 



Art 
acted 
t reat 
rtake 
decom 
ent ra 
i v i n g 
tored 
exam 
ate-r 
ic st 
i ng o 
rdi ng 
sf orm 
i t i o n 
el ate 
, 1 ev 
peat , 
ent ra 
ulfat 
he ef 
, f er 
t i on 
eases 
ciate 
rtant 
and e 



i f i c i 

cons 
ment 
n to 
posed 
t i o n s 

synt 

for 
i ned 
educi 
ate t 
f hig 

to p 
ati on 
s. T 
d wel 
el of 

redu 
t i o n s 
e red 
f 1 uen 
rous 
of su 

in t 
d wi t 

to t 
n v i r o 



al (con 
i d e r a b 1 
of acid 
ex ami ne 
peat o 
, in la 
heti c m 
pH, iro 
for iro 
ng bact 

the e 
hi y vis 
rof i 1 e 

to fer 
he a d d i 

1 with 
sul fat 

c i n g Eh 
of sul 
uct i on 
t conce 
and fer 
lfate t 
hese eh 
h the f 
he long 
nments . 



struct 
e atte 
i c m i n 
the e 
n pH, 
borato 
i n e w a 
n , and 
n , sul 
e r i a . 
x p e r i m 
i b 1 e o 
and ef 
rous i 
ti on o 
parame 
e-redu 
, sulf 
fide o 
s i g n i f 
n t r a t i 
r i c i r 
o the 
emical 
orm at i 
-term 



ed) wetl 
ntion in 
e waters 
ffects o 
and iron 
ry-scal e 
ter. Ef 

sul fate 
fate, su 

Additio 
ental un 
x i d i z e d 
fluent d 
ron unde 
f sulfat 
ters of 
c i n g b a c 
ate loss 
btained . 
icant de 
ons of t 
on, and 
synthet i 

const it 
on of i r 
ret ent io 



and s 
conn 
. A 
f Sph 



and 
reac 
f 1 uen 
. De 
lfide 
n of 
its r 
i ron 
ata , 
r red 
e to 
sul fa 
t e r i a 
, and 
Und 
creas 
otal 
sul fa 
c mi n 
uent s 
on su 
n of 



ystems 
e c t i o n 
study w 
agnum s 
sul fate 
tors 
t was 
pth pro 
, pH, E 
iron in 
esul ted 
which , 
underwe 
u c i n g 
the sys 
te redu 
enrich 

er cond 
es occu 
iron, s 
te foil 
e water 

were 
1 fides 
iron in 



have 
with 
as 
pp. 



files 
h , and 

the 

i n 

nt 

tern 

c t i o n , 
ed in 

i t i o n s 
rred 
ol ubl e 
owing 



the 



INTRODUCTION 



Ipa 
Drai nage 
Conf eren 
Soci ety 
and the 
(Bureau 
Mining R 
17-22. 1 

2 Jo 
A s s i s t a n 
Prof esso 
Env i ronm 
and Dept 
State Un 



per p 

and 
ce sp 
of Su 
U.S. 
of Mi 
eel am 
988 P 
natha 
t and 
r of 
ental 
. of 
i v e r s 



resented 
Surface 
onsored 
rface Mi 
Departme 
nes and 
a t i o n an 
i 1 1 s b u r g 
n M . Die 

Ri chard 
Envi ronm 

Pol luti 
Civil En 
ity, Uni 



at the 1988 Mine 
Mine Reclamation 
by the American 
ning and Reclamation 
nt of the Interior 
Office of Surface 
d Enforcement), April 
h, PA. 
tz is a Graduate 

F . Unz is a 
ental Microbiology, 
on Control Program 
gineering, The Penn 
versity Park, PA. 



of b 
form 
bear 
Such 
a se 
ox id 
iron 
Thio 



Acidic 
oth dee 
ed when 
i n g m i n 

d i s t u r 
r i e s of 
a t i o n - r 
- and s 
bac i 1 1 u 



ox id 



i sti 

meta 

and 

1000 

wate 

from 



ans . A" 
call y c 
1 ions , 
sul fate 

mg/|_. 
rs vary 

pH 2-3 



mine d 
p and s 

water 
eral s a 
bance o 

chemi c 
e d u c t i o 
ul f ur-o 
s ferro 
c i d i c m 
o n t a i n s 

partic 

often 

The mi 
, with 

up to 



rat nag 
urf ace 
and ai 
s s o c i a 
f coal 
al , an 
n reac 
x i d i z i 
o x i d a n 



i ne dr 

high 
ul arly 
at lev 
neral 
pH val 
pH 6. 



e, a was 
coal m i 
r contac 
ted with 
seams s 
d b i o c h e 
tions me 
ng bacte 
£ and T. 
a i n a g e c 
concent r 
iron s p 
el s grea 
acidity 
u e s in t 



te pr 
ning, 
t sul 

the 
t i m u 1 
mical 
di ate 
ri a , 

thi o 
harac 
a t i o n 
ecies 
ter t 
of th 
he ra 



oduct 

i s 
fide- 
coal . 
ates 

d by 
e.g. , 

ter- 
s of 

> 

han 

e 

n g i n g 



310 



Treatment of acidic mine dr 
required in order to meet water q 
standards set by Federal and Stat 
tions. Although costly and labor 
sive, physi cal -chemi cal processes 
been the usual methods of treatme 
Lime, limestone, or sodium hydrox 
employed for neutralization of th 
waters followed by oxidation and 
ation of the metals. More recent 
biologically based systems, inclu 
freshwater wetlands, have been in 
igated for application to the par 
full treatment of acidic mine wat 
and Dietz, 1986). Interest in we 
for acidic mine water renovation 
developed in response to the coal 
industry's need for a more econom 
method of treating acidic mine dr 
Cursory research on volunteer wet 
receiving acidic mine drainage ha 
the belief that such environments 
effect a decrease in the mineral 
uents of the mine waters. These o 
tions prompted laboratory and fie 
investigations aimed at understan 
advantages and limitations of wet 
acidic mine drainage treatment an 
underlying mechanisms responsible 
et al . (1984) studied Sphagnum mo 
in laboratory troughs and found t 
biomass capable of uptake and abs 
of iron and manganese from synthe 
water. Snyder and Aharrah (1984) 
a Typha community existing in a m 
region and observed considerable 
of accumulated iron and manganese 
iated with the plants. Tarleton 
(1984), based on results from con 
microcosms containing Sphagnum mo 
proposed mechanisms of iron remov 
wetland systems which included pi 
uptake, microbial oxidation and 
precipitation, absorption by orga 
deposits, and sulfide precipitati 

Rapid accumulation of incom 
decayed plant material in wetland 
ities results in extensive peat d 
and the formation of anaerobic co 
Decomposition by obligate and fac 
anaerobic bacteria may account fo 
percentage of the reduction of or 
material in a wetland system. In 
to pathways of intramolecular fer 
ation, organic carbon compounds m 
anaerobi cal 1 y respired utilizing 
iron and the sulfur of sulfate as 
electron acceptors. 

Relative to free ferrous 
ferric iron is fairly insolubl 
becoming more so at higher pH 
natural aqueous environments, 
is precipitated as an insolubl 
oxyhydroxide , or jarosite. Ta 
al . (1984) noted that iron oxi 
accumulated within their exper 
units owing to the low solubil 
oxidized iron. The stability 
iron deposits in wetlands is n 
understood. Oxides of iron ar 
microbial reduction when expos 
anaerobic conditions (Nealson, 
addition, ferric iron may be r 
ferrous iron by radical groups 



ai nage l s 
ual i ty 
e regula- 

i n t e n - 

have 
nt. 

i d e are 
e aci d 
p r e c i p i t - 

ly, 

ding 
vest- 
t i al or 
ers ( Unz 
tl ands 
has 

mining 
ical 
ai nage . 
1 ands 
s led to 

may 
c o n s t i t - 
bserva- 
ld 

ding the 
lands for 
d the 

Gerber 
ss placed 
he 

o r p t i o n 
tic mine 

examined 
ine water 
amounts 

assoc- 
et al . 
st ructed 
ss , 
al by 
ant 

ni c 
on . 
pletely 

commun- 
e p o s i t s 
n d i t i o n s . 
u 1 1 a t i v e 
r a large 
g a n i c 

addition 
ment- 
ay be 
f erri c 

f i nal 



l ron , 
e in water 
values. In 
ferric iron 
e oxide , 
rleton et 
des 

i m e n t a 1 
ity of 
of oxidized 
ot well 
e subject to 
ed to 

1983). In 
educed to 

associ ated 



with 
(Szil 
possi 
sol ub 
depos 
treat 
S 
e n v i r 
tions 
reduc 
wetl a 
organ 
conce 
envi r 
s u 1 f i 
reduc 
rich 
syste 
limit 
sul fa 
demon 
sul fa 
ments 
fresh 
(1968 
reduc 
inter 
Herli 
reduc 
recei 
obtai 
sul fa 
chara 
throu 
sol ve 
Tarl e 
sulfi 
al tho 
other 
sul fa 
ions 

exami 

water 

from 

and p 

trans 

of th 

reduc 

incre 

conce 

i n s o 1 



h u m i c m 

agyi, 1 

bil ity 
i 1 i z a t i 
its may 

acid m 
ul fate 
onment 

prevai 
ing env 
nds of 
i c , ana 
ntratio 
onment 
de prod 
ing b a c 
sedimen 
ms, pro 
ed by t 
te pres 
strated 
te redu 

of sul 
water s 
) obtai 
ing bac 
c e p t i n g 
hy and 
t i o n in 
ving ac 
ned by 
te redu 
c t e r i s t 
gh i ncr 
d i ron 
ton et 
des in 
ugh, in 

iron f 
te redu 
in wetl 
The pre 
ne the 
s on 1 a 
t ranspl 
eat, w i 
f ormat i 
e peat 
t i o n in 
a s i n g t 
ntratio 
ubl e su 



aterials and s 
971 and Temple 
exists that ap 
on of the prec 

occur in wetl 
ine drainage, 
reduction occu 
where strong r 
1 . A wel 1 know 
ironment is th 
coastal region 
erobic sedimen 
ns of sulfate 
provide for hi 
uction mediate 
teria. Althou 
ts exist in fr 
duction of sul 
he low concent 
ent. Smith and 

increased pot 
ction with ine 
fate in experi 
ediments. Tut 
ned evidence o 
teria in wood 

acidic mine d 
Mills (1985) o 

the sediments 
i d i c mine d r a i 
these workers 
ction may amel 
ics of acidic 
eased pH and r 
in the form of 
al . (1984) ide 
their experime 

small amounts 
ractions. The 
ction to the b 
ands has not b 
sent study was 
impact of synt 
boratory micro 
anted wetland 
th respect to 
ons within the 
and (2) potent 

the peat with 
he pH and deer 
ns through for 
Ifide deposits 



u 1 f i d e s 

, 1964). The 

p r e c i a b 1 e 

i p i t a t e d mineral 

ands used to 



rs l 
educ 
n su 
e sa 
s . 

t s a 
pres 
gh r 
d by 
gh o 
eshw 
fide 
rati 

Klu 
ent i 
reas 
ment 
tie 
f su 
chi p 
rai n 
bser 

of 
nage 
i n d i 
i ora 
mine 
emov 

i ro 
ntif 
ntal 

com 

imp 
i ndi 
een 

con 
heti 
cosm 
Spha 



(1 

red 
i al 

res 

eas i 
mat i 



n nea 
i ng c 

I fate 

I I ma 
The r 
nd hi 
ent i 
ates 

sulf 
r g a n i 
ater 

i s u 
ons o 
g (19 
al fo 
ed su 
s wit 
et al 
Ifate 

dams 
age a 
ved s 
a 1 ak 
. Re 
cate 
te th 

drai 
al of 
n sul 
i ed i 

micr 
pared 
ortan 
ng of 
resol 
ducte 
c mi n 
s , fa 
gnum 



metal 
u c i n g 
for s 
pect 
ng i r 
on of 



rly any 
o n d i - 

rsh 

i c h 1 y 

gh 

n this 

of 

ate- 

cally 

sual ly 

f 

81) 

r 

pple- 

h 



nd 

ul fate 

e 

suits 

that 

e 

nage 

dis- 
fides, 
ron 
ocosms , 

to 
ce of 

metal 
ved. 
d to 
e 

s h i o n e d 
spp. 

oxide 

zone 
ul fate 
to 
on 



MATERIALS AND METHODS 



Experimental Units 



40 cm 
glass 
ports 
A var 
Harva 
Appar 
del i v 
55-L 
conti 
t u b i n 
betwe 
and s 
i n s e r 
enti r 
envi r 



Three 
x 20 
and 
to p 
i a b 1 e 
rd pe 
atus 
er sy 
stora 
nuous 
g was 
en th 
i 1 i c o 
ted i 
e app 
onmen 



rect 

cm) 
fitte 
rovid 

spee 
ri st a 
Co. , 
nthet 
ge ve 

f 1 ow 

used 
e fee 
n tub 
n the 
aratu 
tal r 



angul 
were 
d wi t 
e a w 
d , mu 
ltic 
Milli 
i c ac 
ssel 
rate 
i n p 
d res 
ing 

peri 
s was 
oom m 



ar chambe 
construct 
h inlet a 
ater dept 
1 1 i c h a n n e 
pump (Har 
s, MA) wa 
i d i c mine 
to the un 

of 2.8 
urn pi ng 
e r v o i r 
(0.125 
s t a 1 1 i c 

1 ocated 
ai ntai ned 



rs ( 
ed o 
nd o 
h of 
1 , m 
vard 
s us 

wat 
its 
L/mi 
nnec 
d th 
ch I 
ump . 
in a 

at 



30 cm x 
f p 1 e x i - 
utlet 

12 cm. . 
odel 1203 

ed to 

er from a 

at a 

n. Tygon 

tions 

e units 

D) was 

The 
n 
20O C and 



311 



Table 1. Composition of synthetic water 
applied to the microcosms during 
experimentation 



Influent Composition(mg/L) 
Stabi 1 i zat i on 
Component Water Stagel Stage2 Stage3 



Fe(III) 

Fe(II) 

S0 4 

CI 1 

Ca 

Mg 

Na 

K 

NH4-N 

NO3-N 

PO4-P 

PH 



1 Chloride concentrations > reported 
value due to pH adjustment with HC1 



60 percent RH. A cool white fluorescent 
light source providing approximately 120 
ft-candles to the surface of microcosms 
was placed on a 12-hour on-off cycle. 






30.0 














30.0 


30.0 











100.0 


26.5 


81.5 


62.5 


1.0 


2.0 


2.0 


2.0 


9.0 


2.0 


2.0 


2.0 


7.5 


5.0 


5.0 


5.0 


5.0 


6.0 


6.0 


6.0 


6.0 


1.5 


1.5 


1.5 


1.5 


0.5 


0.5 


0.5 


0.5 


1.1 


1.1 


1.1 


1.1 


3.5 


3.5 


3.5 


3.5 



Microcosms 



obta 

natu 

The 

remo 

bags 

port 

pi ac 

subs 

the 

i ze 

the 

appr 

evap 

wate 

i zat 

pH 

salt 

was 

micr 

wate 

tati 

cond 

(tre 

vess 

synt 

give 

rece 

iron 

effe 

f err 

Stag 

synt 

(Fe 

unit 

cont 

exi s 

micr 

of t 



Sphag 
i ned 
ral w 
Sphag 



num s 
from 
etl an 
num b 



ved s 

. Th 
ed to 
ed in 
equen 
col 1 e 
under 
stabi 
x i m a 
orati 
r. Af 
ion, 
f 3.5 
s wit 
pumpe 
ocosm 
r for 
on wi 
uct ed 
atmen 
els. 
h e t i c 
n i n 
i ved 
(Fe 
cts 
i c i r 
e two 
h e t i c 
II ) a 
s rec 
a i n e d 
tence 
ocosm 
he sy 



ucces 
e col 

the 

the 
tly, 
c t i n 

cont 
1 i zat 
tely 
on wa 
ter t 
a syn 

with 
hout 
d i nt 
s wer 

a 10 
th th 

i n t 
ts) w 

The 

wate 
table 
synth 

HI) 

f the 

on an 

and 

wate 

nd , i 

e i v e d 

sulf 

of s 

s and 

nthet 



pp. a 
the B 
d i n 
i m a s 
s i v e 1 
1 ecte 
labor 
exper 
were 

area 
rol 1 e 
ion p 
4 wee 
s rep 
he in 
theti 

HC1 
iron 
the 
e mai 
-week 
e thr 
hree 
i t h i n 
chemi 
r use 

1. 
et i c 
in or 

peat 
d i ro 
stage 
r con 
n add 

synt 
ate t 
ulfat 

its 
i c mi 



nd u 
ear 
Cent 
s an 
y an 
d ma 
ator 
imen 
fill 

and 
d co 
erio 
ks , 
1 eni 
i t i a 
c wa 
and 
and 

mic 
ntai 

per 
ee m 
succ 

the 
cal 
d i n 
Stag 
wate 
der 

on 
n ox 

thr 
tain 
itio 
heti 
de 
e re 
effe 
ne d 



nderl 
Meado 
ral P 
d the 
d pi a 
t e r i a 
y and 
tal u 
ed wi 

alio 
n d i t i 
d , wh 
water 
shed 
1 per 
ter, 
cont a 
sul fa 
rocos 
ned 
iod . 
icroc 
e s s i v 

same 
compo 

each 
e one 
r wit 
to ex 
the s 
ide d 
ee un 
i ng r 
n , st 
c wat 
termi 
duct i 
cts 
r a i n a 



ying peat 
ws area, 
ennsy 1 van 

peat wer 
c e d in pi 
Is were t 

immedi at 
nits whic 
th water 
wed to st 
ons. Duri 
i c h 1 a s t e 

loss due 
with d i s t 
iod of st 
ad j usted 
i n i n g m i n 
te (table 
ms. The 
n synthet 

Experime 
osms was 
e stages 

experime 
sit ion of 

stage is 

units 
h oxidize 
amine the 
tabi 1 i ty 
e p s i t s . 
its recei 
educed i r 
age three 
er which 
ne the 
on in the 
n the qua 
ge. Effl 



were 
a 

i a . 
e 

a s t i c 
rans- 
el y 
h, 

from 
abi 1 - 
ng 

d 

to 
illed 
abi 1 - 

to a 
eral 

1), 

i c 
n- 



ntal 
the 



d 

of 

ved 
on 



lity 
uent 



total , 
ferrous 
measure 
Each st 
paramet 
period 
of the 
sion of 
determi 
sul fate 
reduci n 
and pH 
a dista 
effl uen 
the wid 
each de 
were de 
the tan 
each lo 
combi ne 
exper im 
stabi 1 i 
recei vi 
water (t 



soluble 

iron, 
d every 
age was 
ers mon 
of 2 we 
m i c r c 
an exp 
nat i on 
s , sul f 
g bacte 
were me 
nee of 
t wal 1 s 
th and 
pth ave 
termi ne 
ks only 
cat i on 
d into 
ental s 
ze for 
ng the 
able 1) 



(fe 
sulf 

oth 

ope 
i tor 
eks . 
sms 
erim 
of f 
ide , 
ri a 
asur 
1 cm 

and 
the 
rage 
d fo 

wit 
from 
one 
tage 
a pe 
i ni t 



r r i c plus 
ate, and 
er day du 
rated unt 
ed remain 
Depth p 
was made 
ental run 
errous an 

PH, eH, 
counts . 
ed in the 

from inf 

at 10, 2 
values ob 
d. Remain 
r the eff 
h samples 

i n d i v i d u 
average s 
s , tanks 
riod of 3 
ial synth 



f err 
pH we 
ring 
il th 
ed st 
r f i 1 
at th 

and 
d tot 
and s 
Prof i 

micr 
1 uent 
0, an 
t a i n e 
ing p 
1 uent 

coll 
al de 
ampl e 
were 

week 
etic 



ous ) , 
re 

treatm 
e effl 
able f 
e anal 
e cone 
i n c 1 u d 
al i ro 
ul fate 
1 es of 
ocosms 

and 
d 30 c 
d from 
r f i 1 e 

end 
ected 
pths 
. Bet 
al 1 owe 
s whil 
feed 



and 

ent . 

uent 

or a 

ysis 

lu- 

ed a 

n, 

Eh 
at 

m of 

s 
f 
for 

ween 
d to 
e 



Analytical Procedures 

Total and ferrous iron was deter- 
mined using the phenanthrol i ne method 
(APHA, 1975). Ferrous iron was determined 
directly. Total iron was determined after 
acidification of samples and reduction of 
ferric iron with hydroxyl ami ne. Ferric 
iron was determined indirectly by 
subtracting ferrous iron from total iron. 
Sulfate analysis was performed turbid- 
imetrically (APHA, 1975). Sulfide was 
analyzed by the methylene blue method 
(APHA, 1975). The pH and Eh values were 
measured with an Orion 399A pH/mV Analog 
Meter (Orion Research, Cambridge, MA) 
equipped with a Fisher Universal Glass pH 
electrode and a Fisher Platinum Indicating 
E n Electrode (Fisher Scientific Co., 
Pittsburgh, PA). Sul f ate-reduci ng bacteria 
were enumerated using a modification of 
the 5-tube MPN method described by 
Postgate (1984). Serially diluted samples 
were dispensed to 24-well Corning tissue 
culture plates (Fisher Scientific Co., 
Pittsburgh, PA) and incubated at 23°C in 
an anaerobic chamber under an atmosphere 
of N 2 . 



Data Analysis and Statistical Procedures 

Effluent results of stage 2 and stage 
3 units were statistically compared to 
evaluate the effects of sulfate reduction 
on iron and pH. Data collected during the 
last 2 weeks of stage two and stage three 
were averaged and statistically compared 
using a one-tailed AN0VA of comparison of 
means (Sokal and Rohlf, 1981). 



RESULTS AND DISCUSSION 

Ferric Iron Reduction 

In the first stage of experimental 
work, microcosms, which had been purged of 



312 



4.50 



4.25 



^ 4.00 



3.76 



12.0 



o pH o Ferrous Iron 

a Total Iron 



y 



/ 



Oc^*v 



'^y 







3 gn i 1 1 1 1 1 1 1 1 1 1 1 1- 

I 2 3 4 5 6 7 8 10 II 12 13 14 



Tlme(weeks) 



5.60 



5.00 



4.50- 



4.00- 



3.50r 



3.00 



o Influent pH a Effluent pH 
a Influent Eh or Effluent Eh 



B 




030 
500 

■ 1400 
-300 

-200 

TlOD 

a 



£ 



12 3 4 6 8 7 

DepthCcm) 

Fig.l. A. Effluent quality of stage one 
microcosms. B. Depth profile of 
influent and effluent pH and Eh 
parameters of stage one microcosms. 
Values are the mean of individual 
measurements on three microcosms. 



•100 



all 

synt 

oxid 

i ron 

resu 

v i s i 

Subs 

the 

cons 

of f 

depo 

1 eve 

(fig 

of a 

depo 

prof 

run 

infl 

mark 

1 owe 

Stag 

(fig 

dimi 
f err 
rel a 

mi cr 



trace 
heti c 
i zed 
(Fe(I 
lted 
ble i 
equen 
efflu 
idere 
e r r i c 
sits. 
Is, b 
. 1A) 
grea 
sited 
ile d 
are p 
uent 
edl y 
r at 
e one 
. 5) 
n i s h e 
i c i r 
t i v e 1 
ocosm 



s of f 

water 
form. 
D) to 
in the 
ron ox 
t appe 
ent of 
d evid 

iron 

The 

oth to 

may r 
ter pr 

i n th 
ata ob 
resent 
and ef 
depres 
the ef 

f erro 
i ncrea 
d with 
on red 
y high 
s , sug 



errous 

conta 

The c 

the e 

forma 

i d e d e 

arance 

the m 

ence o 

and th 

i n c r e a 

tal an 

ef 1 ect 

oport i 

e redu 

tai ned 

ed in 

f 1 uent 

s e d w i 

f 1 uent 

us i r o 

sed si 

depth 

u c t i o n 

E n va 

g e s t i n 



i ron , r 
i n i n g i r 
h a r g i n g 
x peri men 
tion of 
posits i 

of ferr 
i c r o c o s m 
f intern 
e iron o 
sed effl 
d Fe(II) 

the tra 
on of th 
cing zon 

at the 
fig. IB. 

redox p 
th depth 

regi on 
n concen 
i ghtl y a 
. Intere 

occurre 
lues wit 
g that i 



ecei 
on i 

of f 
tal 
high 
n th 
ous 
s wa 
al r 
xide 
uent 
, ov 
nsf o 
e i r 
e . 
end 

Bo 
oten 
and 
of t 
t rat 
s E n 
st i n 
d at 
hi n 
ron 



ved 
n the 
e r r i c 
units 

iy 

e peat. 

iron in 

s 

eduction 

iron 
er time 
rmat i on 
on oxides 
Depth 
of the 
th 
t i al were 

were 
he units . 
ions 

val ues 

giy. 

the 
oxide 



Table 2. Depth profile counts of sulfate 
reducing bacteria of stage two and 
stage three. 



Sulfate-Reducing Bacteria(counts/mL) 
Stage 2 Stage 3 

Tank 12 3 12 3 



68000 3600 6800 

6800 4400 1100 

1800 28000 8800 

36000 25000 36 



deposits may be unstable towards solubil- 
ization and transformation within a major 
segment of the peat column. 



Depth 








(cm) 








2 


5.6 


36 


360 


4 


4.4 


68 


68 


7 


880 


3200 


680 


10 


56 


5200 


104 



Sulfate Reduction 



Th 
a 1 ow 
ex ami n 
study, 
wetl an 
val ues 
acti vi 
there 
values 
not re 
pH (Tu 
presen 
microb 
depres 
mi croc 
reduci 
all de 
n e g 1 i g 
(<0.5m 
precl u 
Nevert 
reduci 
units 
sul fat 
microc 
c o n d i t 
throug 
sul fat 
s i g n i f 
( as mu 
infl ue 
were a 
Depth 
depict 
s u 1 f i d 
d i s s i m 
p r i n c i 
Decrea 
depth 
cosms 
possi b 
sul fat 
to dif 
rather 
dispel 
effl ue 
p r o f i 1 
bal anc 



e pot 
PH, f 
ed in 
Mic 
d env 

i s n 
ty an 
is us 

repr 
fleet 
ttle 
ted a 
i al s 
sed p 
osms . 
ng ba 
pths 
ible 
9/L) 
ded s 
hel es 
ng ba 
in sp 
e. Th 
osm ( 
ions 
hout 
e was 
i cant 
ch as 
nt an 
c h i e v 
p r o f i 
i ng m 
e (fi 
i 1 a t o 
pal p 
sing 
were 
d u r i n 
ility 
e con 
f u s i o 

than 
led i 
nt su 
es of 
e of 



ent i a 
reshw 

stag 
robi a 
i ronm 
ot CO 
d whe 
ual ly 
esent 
i ve o 
et al 
ppear 
ulfat 
H val 

High 
c t e r i 
of mi 
infl u 
i n st 
i g n i f 
s, hi 
c t e r i 
i te o 
e pro 
fig. 
were 
the s 

1 i m i 

redu 

10 m 
d eff 
ed by 
1 es o 
easur 
g. 4) 
ry su 
roces 
sul fa 
obser 
g sta 

that 
cent r 
n 1 im 

sulf 
n con 
lfate 

cons 
sul f i 



1 for 
ater 
es tw 
1 sul 
ents 
n s i d e 
n sue 

the 

bulk 
f the 
., 19 

to s 
e-red 
ues i 

numb 
a ( ta 
croco 
ent s 
age t 
i c a n t 
gh 1 e 
a exi 
f the 
file 
2B) r 
event 
e d i m e 
ted. 
ct i on 

g so 4 

1 uent 

stag 
n sta 
ement 

and 
lfate 
s of 
te co 
ved i 
ge th 

the 
at ion 
it at i 
ate t 
s i d e r 

cone 
ervat 
de pr 



sul fate 
envi ronme 
o and thr 
fate redu 
at decide 
red to be 
h an even 
suspicion 

water me 

sediment 
68). The 
upport th 
ucing act 
n the exp 
ers of su 
ble 2) we 
s m s s t u d i 
ul fate co 
wo of the 

sulfide 
vels of s 
sted in s 

absence 
of the st 
eveal s th 
ual ly est 
nt even t 

In add it 
s of effl 
- 2 /L by c 

[fig. 3A 
e three m 
ge three 
s of sulf 

Eh (fig- 

reducti o 
sulfate r 
ncentrati 
n all thr 
ree exper 
gradual ly 
s may hav 
o n s in t h 
ransf orma 
ation of 
entration 
i v e ions, 
oduced ag 



reduct 
nt was 
ee of 
c t i o n 
dly ac 

a ma j 
t occu 

that 
asurem 

micro 

resul 
e prem 
ivity 
erimen 
Ifate- 
re f ou 
ed , ho 
ncent r 

study 
format 
ul fate 
tage t 
of exo 
age tw 
at red 
a b 1 i s h 
hough 
ion, 
uent s 
ompari 
] val u 
i c r o c o 
microc 
ate a n 
3) poi 
n as t 
emoval 
ons wi 
ee mi c 
i m e n t s 

dimi n 
e been 
e s e d i 
t i ons 
the co 
s, dep 

and t 
ai nst 



ion in 

the 

i n 

id pH 

or 

rs 

the pH 

ents 

zonal 

ts 

i s e of 

at 

tal 

nd at 
wever , 
a t i o n s 

ion . 

wo 

genous 

o 

ucing 

ed 



ul fate 

son of 

es) 

sms . 

osms 

d 

nt to 

he 

th 

ro- 

. The 
i shed 

due 
ment s 
was 

nstant 
th 
he 



313 



4.00 



4 * pH ° Ferrous Iron 

a To ta I I ran 




R 



30.0 



•24.0 



,^^-1 



■s*s4 



3.00 



18.0 



12.0 



0.0 



8 1! IB 21 28 31 38 41 48 61 
TlmeCdays) 



0.0 



20.0 



10.0 






e.o 



if 



4.0 



0.0 



4.00 



3.80 



■ 3.20 



3.00 



o pH d Total Iron 
a S04 * Ferrous Iron 



te**-^^ 



n 



120.0 




II 16 



21 28 31 38 
Tlme(daye) 



100.0 



41 48 61 68 



0.0 



4.00 



3.00 



o Influent pH ° Effluent pH D 

a Influent Eh ■ Effluent Eh L ~ J 



i7DD 



000 




e 



Fig 



two 



2 3 4 6 

Depth(cm) 

2. A. Effluent quality of stage 
microcosms. B. Depth profile of 
influent and effluent pH and Eh 
parameters of stage two microcosms 
Values are the mean of individual 
measurements on three microcosms. 



-100 



x 
a 



4.00 



3.76 



3.60 



3.25 



3.00 



o Influent pH 

a Influent Eh 



a Effluent pH 
■ Effluent Eh 




ir200 



1 



6 



-300 



2 3 4 6 
Depth(cm) 
Fig. 3. A. Effluent quality of stage three 
microcosms. B. Depth profile of 
influent and effluent pH and Eh 
parameters of stage three microcosms 
Values are the mean of individual 
measurements on three microcosms. 



sul fate 
produce 
have be 
to thes 
was pro 
units, 
and 3B) 
two and 
val ues 
r e d u c i n 
reduct i 
sulfide 
values 
likely 
of o x i d 
Sed 
r e d u c t i 
i ncreas 
hydroge 
p r e c i p i 
effect 
iron co 
stat i st 
resul t s 
stage t 
stage t 



lost, 
d by st 
en d e r i 
e units 
d u c e d i 
Compari 

in the 

three 
were mo 
g c o n d i 
on woul 

woul d 
found i 
to be a 
ized fo 
iment s 
on mi gh 
e in p H 
n ions 
tate fe 
of sulf 
ncent ra 
i c a 1 1 y 

(figs. 
wo (min 
hree (m 



The 1 
age t 
ved f 

sine 
n sul 
sons 

mi cr 
revea 
re ch 
t i o n s 
d be 
be pr 
n sta 
s a r 
rm of 
d i s p 1 
t be 

owi n 
by su 
rrous 
ate r 
t i o n s 
compa 

2A a 
or su 
ajor 



arge 
h ree 
rom t 
e no 
fate- 
Of E h 
ocosm 
led s 
aract 

expe 
occur 
esent 
ge th 
esul t 

sulf 
aying 
expec 
g to 
lfide 

iron 
educt 

and 
ring 
nd 3A 
lfate 
sul fa 



amounts 
microcos 
he sul fa 
detectab 
1 i mi ted 

val ues ( 
s during 
tage thr 
eristic 
cted whe 
ring and 
. The lo 
ree unit 

of the 
ate to s 

active 
ted to e 
consumpt 

ions an 

as FeS. 
ion on e 
pH was a 
the effl 
) obtain 

reduct i 
te reduc 



of sul 
ms mus 
te cha 
1 e sul 
stage 
figs. 

stage 
ee E h 
of the 
re sul 

in wh 
wer Eh 
s are 
reduct 
u 1 f i d e 
sul fat 
ffect 
ion of 
d to 

The 
f f 1 uen 
nalyze 
uent 
ed wit 
on ) an 
t i o n ) 



fide 

t 

rged 

fide 

two 

2B 

s 



fate 
i ch 



i on 



e 
an 



t 
d by 



system 
iron i 
there 
effl ue 
to sta 
in pH 

Ef 
forms) 
in sta 
two un 
effl ue 
as a r 
stage 
1 ow Eh 
concen 
peat . 
concen 
three 
of sta 
preci p 
Al thou 

(fig. 

s i m i 1 a 
depos i 
microc 
preci p 



s ; bo 
nf 1 ue 
to be 
nt i r 
ge 3, 
coul d 
f 1 uen 
were 
ge th 
its. 
nt fe 
esul t 
3 exp 

(fig 

t r a t i 
Lower 
t r a t i 

(fig. 

ge tw 
i t a t i 
gh de 
5) in 
r the 
ts in 
osms 
i tati 



th o 
nt . 

sig 
on c 

but 

be 
t ir 

1 ow 
ree 

Hig 
rrou 

of 
erim 
. 3B 
ons 

eff 
ons 

3A) 
o (f 
on o 
pth 

sta 

app 

the 
was 
on. 



f whi 
The 
n i f i c 
oncen 
no s 
estab 
on co 
er by 
units 
h con 
s i ro 
i n c r e 
ental 
) val 

(fig. 

1 uent 
actua 

comp 
ig. 2 
f the 
prof i 
ge tw 
earan 

peat 
sugge 

Acid 



ch re 
i n v e s 
ant d 
t r a t i 
i g n i f 
1 i she 
ncent 

at 1 

as c 
cent r 
n wou 
ased 

unit 
ues a 

4) P 

f err 
lly r 
ared 
A) pr 

i ron 
1 es o 
o and 
ce of 

of s 
stive 

dige 



cei ve 
t i g a t 
ecrea 
ons f 
i c a n t 
d. 

ratio 
east 
ompar 
at ion 
Id be 
reduc 
s as 
nd hi 
resen 
ous i 
esul t 
to th 
esuma 

by s 
f fer 

thre 

vi si 
tage 

of i 
stion 



d a ferrous 
ion found 
ses in all 
rom stage 2 
i ncreases 



ns ( a 
50 pe 
ed to 
s of 

expe 
t i o n , 
a res 
gh su 
t i n 
ron 
ed i n 
e eff 
bly d 
ulfid 
rous 
e are 
ble b 
three 
ron s 

of t 



11 

rcent 
stage 

cted , 

i n 
ult of 
lfide 
the 

stage 
1 uent 
ue to 
e. 
i ron 

1 ack 

ul fide 
he 



314 



120 



30.0 



o Sulfate 



a Sulfide 



24.0 




DepthCcnO 

Fig. 4. Depth profile of sulfate and 

sulfide in stage three microcosms at 
the conclusion of experimental run. 
Values are the mean of individual 
measurements on three microcosms. 
40.0 1 



X.O 



-sao.o 



10.0 



O Stage I 



A Stage 2 



o Stage 3 





0.0 ■ 



4 8 15 

DepthCcnO 

Fig. 5. Depth profile of ferrous iron for 
all stages of experiment as determined 
at conclusion of experimental run. 
Values are the mean of individual 
measurements on three microcosms. 



deposits 
i ndi cated 
possi bl y 
i ron sul f 
was i n t e n 
three m i c 
10 to 20% 
became ox 
holding t 
oxidized 
readi 1 y s 
sediments 
whereas , 
remain in 
As no 
erence be 
two and s 
over time 
might be 
hydrogen 
cosms, th 
owi ng to 



f ol 1 owed 
the pre 
as a res 
ides . A 
ded for 
rocosms , 
of the 
i d i z e d d 
ank to t 
iron wou 
o 1 u b i 1 i z 

of the 
FeS prec 
tact . 
ted earl 
tween ef 
tage th r 
. Altho 
expected 
ion cone 
is did n 
pref eren 



by ferrous iron analysis 
sence of ferrous iron 
ult of the breakdown of 
Ithough only ferrous iron 
delivery to stage two and 

in reality, an estimated 
iron in the influent 
uring transport from the 
he microcosms. The 
Id be expected to become 
ed in the reducing 
stage three microcosm, 
ipitates presumably would 

ier, no significant diff- 
fluent pH values of stage 
ee microcosms was found 
ugh sulfate reduction 

to produce lower 
entrations in the micro- 
ot occur, presumably 
tial reaction of the 



Table 3. Statistical comparison of stage 2 
and stage 3 results. Averages of the 
three experimental units. 

Statistical Comparison 
Averages(mg/L) 
Stage2 Stage3 Probabi 1 i ty 



Total Fe 

Fe(II) 

pH 

Sul fate 



18.5 

8.74 

3.3 



10.1 

2.69 

3.2 

91.4 



<0.001 
<0.001 

1 



1 Hypothesis rejected ( Stage2 > Stage3) 



sulfide 
hydrogen 
present 
iron may 
consumpt 
ef f 1 uent 
of stage 
that sul 
iron rem 
excess o 
the pH p 
three un 
reductio 
than tha 
sul fate 
the pH w 
sul fate 
i ncrease 
sul fate 
e x p e r i m e 
ence of 
high i ro 
and the 
Further 
determi n 
peat env 
acid m i n 



ions w 
ions, 
i n exc 
be av 
ion . 
iron 
three 
fate 1 
oval . 
f sulf 
r o f i 1 e 
i ts ) w 
n was 
t of s 
reduct 
as not 
reduct 
d pH m 
reduct 
nt may 
p H imp 
n cone 
shal 1 o 
invest 
e whet 
i ronme 
e wate 



ith f 
Hen 
ess o 
a i 1 a b 
Exami 
and s 

unit 
osses 

In t 
ide w 

of t 
hich 
not s 
tage 
ion t 

incr 
ion i 
ay st 
ion . 

have 
rovem 
entra 
w dep 
i g a t i 
her s 
nt ca 
r . 



errous 
ce , on 
f that 
1 e for 
nation 
ul fate 
s cl ea 
did n 
he pre 
as nev 
he m i c 
e x h i b i 
i g n i f i 
two wh 
ook pi 
eased 
n this 
ill be 
The d 
preve 
ent du 
t i o n s 
th of 
ons ar 
ul fate 
n impr 



i ron 

ly sul 

preci 

hydro 

of in 

conce 

rly de 

ot exc 

sent s 

er obs 

rocosm 

ted su 

cantly 

ere on 

ace. 

as a r 

exper 

a ben 

e s i g n 

nted t 

e to 1 

of the 

the ta 

e need 

reduc 

ove th 



rath 

fide 

pita 

gen 

flue 

ntra 

mons 

eed 

tudy 

erve 

(st 
lfat 

dif 
1 y m 
Al th 
esul 
imen 
efit 
of t 
he o 
ow p 

inf 
nk. 
ed t 
ti on 
e pH 



er than 

ted by 

ion 

nt and 

t i o n s 

trated 

ferrous 

, an 

d and 

age 

e 

f erent 

i nor 

ough 

t of 

t, 

of 
hi s 

ccur r- 
H and 
1 uent 



l n 
of 



The res 
support the 
deposited i 
subject to 
sol ubi 1 i zat 
stances, th 
in organ ica 
a serious c 
developed f 
treatment . 

Sulfate 
the low pH 
experimenta 
influent su 
reduced. Th 
sulfate red 
on ef f 1 uent 
more than 5 
occurred in 
microcosms 
units. The 
the form of 

The imp 
study for t 
earmarked f 
treatment 1 
such as Sph 



CONCLUSIONS 

ults of the laboratory study 

contention that ferric oxides 
n Sphagnum microcosms are 
microbial reduction and 
ion. Under these circum- 
e fate of iron oxide deposits 
lly rich sediments must become 
onsideration in wetlands 
or acid mine drainage 



reduc 
of the 
1 unit 
lfate 
e sul f 
ucti on 

i ron 
0% gre 

the s 
as com 
iron r 

iron 
ortant 
he dev 
or aci 
i e wit 
agnum , 



t i on was 

peat mat 
s with as 
concent ra 
ide produ 

had a si 
concent ra 
a t e r iron 
ul fate-re 
pared to 
emoved wa 
sulfides. 

i m p 1 i c a t 
el opment 
die mine 
h peat-pr 

Typha , C 



demonstrated at 
eri al of the 

much as 10% of 
t i o n s being 
ced through 
gnificant effect 
tions in that 

removal 
d u c i n g 

sul f ate-1 imi ted 
s retained in 

ions of the 
of wetlands 
drainage 
oducing plants, 
arex, and 



315 



S c i r p u s . An artificially constructed 
system, in which the reducing zone is 
sufficient to solubilize and reduce 
oxidized iron species, but lacks the 
presence of sulfide for precipitation of 
the iron, may eventually release soluble 
iron at levels in equilibrium with that of 
the influent. Construction of deep cells 
for the collection of fermentable plant 
matter should assure adequate reducing 
potential for microbial sulfate reduction 
in the wetland bed since acidic mine 
waters normally contain high sulfate 
concentration environment. 



LITERATURE CITED 

1. Amer. Public Health Assoc. 1975. 
Standard methods for the examination of 
water and wastewater. 15th Edition, 
1134p.. APHA. New York, NY. 

2. Gerber, D.W., J.E. Burn's, and 

R.W. Stone. 1985. Removal of dissolved 
iron and manganese ions by a Sphagnum 
moss. p. 365-373. J_n Wetlands and water 
management on mined lands. The Penn- 
sylvania State University, University 
Park, PA. 



3. Herlihy, A.T. and A.L. Mills. 1985. 
Sulfate reduction in freshwater 
sediments receiving acid mine drainage. 
Appl . and Env. Microbiol. 49:179-186 

4. Kennedy, J.L. and R.C. Wil moth . 1976 . 
Combination limestone treatment of acid 
mine drainage, p. 64-99. In The sixth 
symposium on coal mine drainage 
research. Louisville, KY. 

5. Mitsh, B. and J. 6. Gosselink. 1986. 
Wetlands. 539 p. Van Nostrand Reinhold 
Co., New York, NY. 

6. Nealson, K.H. 1983. The microbial iron 
cycle. In Microbial geochemistry. W.E. 
Krumbei rTfed . ) . Blackwell Scientific 
Publ . , Boston , MA. 

7. Postgate, J.R.. 1984. The sulphate- 
reducing bacteria. 2nd Edition, 208 
p.. Cambridge University Press, New 
York, NY. 

8. Smith, R.L. and M.J. Klug. 1981. 
Reduction of sulfur compounds in the 
sediments of a eutrophic lake basin. 
Appl. and Env. Microbiol. 41:1230-38 

9. Snyder, CD. and E.C. Aharrah . 1984 . The 
influence of the Typha community on 
mine drainage, p. 149-53. In The 1984 
symposium on surface mine Hydrology, 
sedimentol ogy , and reclamation. 
University of Kentucky, Lexington, KY. 

10. Sokal, R.P. and F.J. Rohlf. 1981. 
Biometry. 2nd Edition, 859 p.. W.H. 
Freeman and Company, San Francisco. 



11. Sparling, H. and B.M. Hennick. 1974. 
The production of hydrogen sulfide in 
peats. Can. J. Microbiol. 19:59-66 

12. Szilagyi, M.. 1971. Reduction of 
Fe ( III) ion by humic acid prepar- 
ations. Soil Sci. 111:233-235 

13. Tarleton, A.L., G.E. Lang, R.K. 
Wieder. 1984. Removal of iron from 
acid mine drainage by Sphagnum 
peat: results from experimental 
laboratory microcosms, p. 413-20. 

In The 1984 symposium on surface mine 
Hydrology, sedimentol ogy , and reclam- 
ation. University of Kentucky, Lexi- 
ngton, KY. 

14. Temple, K.L.. 1964. Syngenesis of 
sulfide ores: Evaluation of biochem- 
ical aspects. Econ. Geol . , 50:1473-91 

15. Tuttle, J.H., P.R. Dugan, C.B. 
MacMillan, C.I. Randies. 1968. 
Microbial di ssimi 1 atory sulfur cycle 
in acid mine water. J. of Bact., 
97:594-602 

16. Unz, R. F. and J. M. Dietz. 1986. 
Biological applications in the treat- 
ment of acidic mine drainages. 
Biotechnol. Bioeng. Symp.,16: 163- 
170. Examination of Water and 
Wastewater" 14th Edit.. Amer. Pub. 
Health Assn., New York. 



316 



IRON AND MANGANESE REMOVAL IN A TYPHA -DOMINATED WETLAND DURING 
TEN MONTHS FOLLOWING ITS CONSTRUCTION 1 



Lisa L. Stillings, Jeffrey J. Gryta, Tedd A. Ronning : 



Abstract. A Typ h a ^ -dominated wetland was 
constructed in September 1986 to remove iron and 
manganese from a surface mine seep characterized by 
an average flow of 10 gpm, pH of 5.5, and maximum 
concentrations of 40, 50, and 2700 ppm Fe, Mn, and 



SO 



2- 



respectively. Typha were planted at l-ft 



centers in a 110 by 20 ft basin containing a basal 
6-inch layer of agricultural-grade limestone and an 
upper 10-inch layer of an equally proportioned 
mixture of peat, compost, and sandy soil. Surface 
waterflow volume was recorded with inlet and outlet 
weirs, and rainfall was measured by a continuous 
recording rain gauge. Water samples were collected 
at the inlet and outlet of the wetland, and at 3 3 
surface locations within the basin. Substrate 
interstitial water was sampled from six, two-level 
piezometer nests. All samples were analyzed for Ca, 
Fe, K, Mg, Na, Mn, P, and S0 4 2 ~, on a biweekly basis 
from October 1986 to July 1987. Seventy-four percent 
of the 18 6 kg of iron introduced to the system was 
removed during this period. Manganese removal was 
less successful; only 8% of the 368 kg influx was 
removed. Periods of increased waterflows 
corresponded with the smallest percent reduction of 
influent mass concentrations. The greatest reduction 
of influent mass occurred during the June 1 sample 
period, when 26.4 kg of iron, and 14.2 kg of 
manganese, were removed from solution. The wetland' s 
ability to remove iron and manganese varied with time 
and location within the basin. Inlet flow volume, 
basin length, and water flow through the substrate 
were important factors affecting the retention of 
influent iron and manganese. 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation 
Conference sponsored by the American 
Society for Surface Mining and 
Reclamation and the U.S. Department of 
the Interior (Bureau of Mines and Office 
of Surface Mining Reclamation and 
Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

Lisa L. Stillings is a graduate 
student, Jeffrey J. Gryta is an Assistant 
Professor, and Tedd A. Ronning is a 
graduate student, Department of Geology, 
Kent State University, Kent, OH. 



INTRODUCTION 

Wetlands have been used since the 
1970 's for the treatment of acid mine 
drainage (AMD) and municipal wastes. 
Studies conducted by Huntsman et al. 
(1978) and Wieder and Lang (1982) have 
compared influent and effluent 
concentrations of AMD passing through 
wetlands. Wieder et al. (1985) and Girts 
and Kleinmann (1986) have examined the 
wetland flow conditions which surround 
AMD amelioration. Recommendations 
regarding the most effective flow rates, 
basin size, and vegetation types have 



317 



been made by Kleinmann et al. (1986) 
Girts and Kleinmann (1986) and Girts et 
al. (1987). Research has also been 
conducted with wetlands for the treatment 
of municipal waste (Wile et al. 1985, 
Kadlec 1987) , and while AMD and municipal 
waste influent possess inherently- 
different qualities, criteria such as 
length and retention time are found to be 
important in the successful treatment of 
these two types of contamination. Detail 
is lacking, however, pertaining to where 
metal removal takes place within 
wetlands, what chemical conditions 
surround removal, and how removal rates 
vary with time. In this study, an 
instrumented wetland was constructed to 
quantify iron and manganese removal over 
a 10-month period following wetland 
construction. The wetland received flow 
from an AMD seep emerging from spoil 
reclaimed following mining of the Middle 
Kittanning (#6) and Lower Kittanning (#5) 
Coals of the Allegheny Formation in 
Tuscarawas County, OH. 

METHODS 

Wetland Construction and Instrumentation 

Based upon an average flow estimate 
of 10 gpm, a 110 by 20 ft basin was 
constructed (fig. 1) following the 
recommended size of 200 ft . of wetland/1 
gpm of flow (Kleinmann et al. 1986). The 
basin was leveled and filled with six 
inches of agricultural-grade limestone 
which, in turn, was covered with a 10- 
inch humic layer consisting of an equally 
proportioned mixture of peat, compost, 
and sandy soil. The limestone was placed 
as a distinct layer and not mixed with 
the humic soil to avoid coating the 
limestone with a ferric hydroxide 
precipitate (Kleinmann et al., 1986). 
Cattails ( Typha ) were collected from the 
surrounding area and planted in the basin 
at 1-ft centers. Cores taken one month 
later showed that these layers had 
compacted to approximately 5.5 inches of 
limestone, and 7 inches of humic 
material . 

The water budget is recognized as a 
key factor affecting water quality and 
wetland functions (Kadlec 1987) . 
Therefore, 40° V-notch weirs with 
continuous water level recorders were 
installed at the inlet and outlet to 
measure surface waterflow into and from 
the wetland. Flow volumes were 
calculated from the following equation 
for sharp-edged, V-shaped weirs 
(Brakersiek et al. 1979): 

Q = C x 8/15 x /2g x tan(6/2) x H 2 " 5 (1) 

where, 

Q = discharge (ft /sec) 

C = weir coefficient (0.582 for a 40° V- 

notch weir) 
6 = total angle of notch (degrees) 
H = head above the lowest point of the 

notch (ft) , and 
g = gravitational acceleration (32.17 

ft 2 /sec) . 



A weighing-bucket continuous recording 
rain gauge was placed at the site to 
measure rainfall; potential 
evapotranspiration was estimated from 
evaporation pan data collected during the 
spring/summer months. Additionally, six, 
two-level piezometer nests were placed 
within the basin: the lower level 
sampled subsurface water within the 
bottom 3 inches of the humic substrate, 
and the upper level sampled the top 3 
inches of the substrate. 

Sampling Procedure and Chemical Analysis 

Wetland water samples were collected 
on a biweekly basis from October 11, 1986 
through July 3, 1987 from the 
piezometers, the inlet and outlet weir 
pools, and from 33 surface sites defined 
by an 11-column by 3-row sample grid 
(fig. 1) . Nineteen separate collections 
were made. Frozen wetland surface 
conditions, and site inaccessibility due 
to excessive rainfall precluded the 
collection of some samples. Chemical 
analysis included Ca, Fe, K, Mg, Na, Mn, 
S0 4 2 ~ and orthophosphorus. Inductively 
Coupled Plasma (ICP) spectrometry was 
used for all analyses except for S0 4 2 ~ 
and ortho-P. These were determined by 
colorimetric procedures outlined by Hach 
(1986) , with one modification: rather 
than reading sample concentration 
directly from the spectrophotometer, a 
series of standards was analyzed with 
each sample batch to establish a 
regression relation of concentration 
versus absorbance. Sample concentrations 
were then determined with this equation. 

Water Budget Analysis 

Water budget volumes were calculated 
as described in LaBaugh and Winter, 1984, 
by dividing the study period into 
intervals whose midpoints correspond to 
the date of water sampling. Budget 
components include inlet flow (If) , 
outlet flow (Of) , rainfall (Ppt) , 
potential evapotranspiration (Pet) , and 
change in volume of surface water stored 
(AS) within the basin (volume determined 
by multiplying wetland area times average 
surface water depth) . All of the above 
components were calculated for each 
interval, and the following equation was 
used to determine the residual (Res) 
component: 



If + Ppt - Of - Pet - Z\S = Res 



(2) 



A positive residual value denotes a net 
water gain or an ungauged loss of water 
(such as vertical seepage down through 
the basin) , while a negative residual 
value denotes a net water loss or an 
ungauged source of water (such as 
vertical seepage up into the basin) . 
Residual values might also reflect 
operator and/or instrument error. 

Mass Budget Analysis 

Following water budget analysis, a 



318 



INLET 




-VVO^WX^N HUMIC LAYER 



WATER SAMPLE LOCATION 

(A,B) ROW(A) AND COLUMN(B) 

DESIGNATION OF SELECTED 
WATER SAMPLE LOCATIONS 



PIEZOMETER LOCATION 
SURFACE WATER FLOW 

HAY BALES 

SURFACE WATER LEVEL 



OUTLET 



LIMESTONE LAYER 



3 FEET 



Figure 1. — Schematic drawing of the wetland basin with sample locations. The insert 
locates the wetland with respect to the AMD seep and the site of pre-wetland 
chemical treatment. 



total mass flux for each interval was 
calculated by multiplying inlet and 
outlet flow volumes by their respective 
inlet and outlet concentrations. 
Assuming that negative residual volumes 
(equation 2) resulted from an ungauged 
source, these were multiplied by the 
inlet concentrations and added to the 
amount of mass entering the wetland. 
Positive residuals were attributed to 
ungauged outflows, and these were 
multiplied by the outlet mass 
concentrations and added to the amount 
leaving the wetland. Total influx was 
calculated as the sum of the influx 
values for each interval. Total outflux 
was determined in a similar manner. 
Inflow and outflow were the only water 
budget components used in mass budget 
calculations because iron and manganese 
concentrations in rainfall were assumed 
to be negligible, and because change in 
surface storage accounted for less than 
one percent of the total water budget. 



RESULTS & DISCUSSION 

Concentrations of Fe, Mn, and SO* 2- 
at the seep varied throughout the study 
period, and while their maximum 
concentrations were 44, 50, and 2720 ppm, 
respectively, they averaged 31 ppm Fe, 3 4 
ppm Mn, and 1130 ppm S0 4 2 ~. Mass 
concentrations at the inlet to the 
wetland averaged 13 ppm Fe, 3 ppm Mn, 
and 990 ppm S0 4 2- ; evidently Fe and 
S0 4 2 ~ were removed from solution while 
the water flowed through a ditch leading 



from the seep to the wetland (insert, 
fig. 1) . The following discussion 
considers only the mass concentrations 
which actually entered the wetland. 

Overall, 74% of the 186.4 kg of iron 
delivered to the wetland during the study 
period remained within the basin. Each 
sample interval within this period 
consistently displayed a smaller mass 
outflux than the influx, indicating that 
iron removal continued through time (fig. 
2a) . The smallest percentage of iron 
removal, 43%, occurred on two separate 
occasions: first, during the December 8 
interval, when 2.4 kg of iron were 
removed from the flow; and second, during 
the April 15 period, when 6.2 kg were 
removed (fig. 2a) . The largest reduction 
occurred over the June 1 sample interval, 
with the removal of 26.4 kg of iron (94% 
of the inlet mass flux) . 

Manganese was not as successfully 
treated by the wetland: only 8.3% of the 
3 68.5 kg of mass which entered the basin 
during the entire study period was 
removed from the flow. During most of 
this period (October 11 through March 18, 
fig. 2b) inlet and outlet manganese flux 
values barely differed except for the 
December 8 and January 5 sample intervals 
when the outlet flow carried an 
additional 5 and 2 kg, respectively. Not 
until mid-May (fig. 2b) did the system 
begin to consistently remove manganese 
from solution, and, as with iron, the 
June 1 period displayed the greatest 
manganese removal, retaining 14.2 kg (26% 



319 



CO 

CO 
D 



A. 

60 -, 



M Inlet 
D Outlet 
Residual 



50 





40 


CO 


30 


CO 




O 


20 




10- 



co i- co ro 

M " J *" J 



60 
50 H 
40 
30 
20 
10 




Sample Date 



Mn 



m 








oo 


CO 




r- 


CD 




CN 


IT) 


O 


CN 


o 


CN 




CM 




CM 




CM 




CN 




r 


vl 




[ 


D 




J 




F 





CN 



ll 



CD 



lO 



M 



r- CO 

M " 



ro 



Sample Date 



Figure 2. --Inlet and outlet components of the iron (A) and manganese (B) budgets 
for each sample period. 



calculated 



of the inlet mass concentration) within 
the wetland. Later, during the July 3 
sample period, inlet manganese 
concentrations were reduced by 40%, a 
percentage corresponding to the removal 
of 13.8 kg. 

Flow, as suggested by Kadlec (1987) , 
does appear to influence metal removal in 
a wetland system. The peak inlet flow 
volume, 2 0.1 gpm, occurred over the 
period of April 15 (figs. 3a, c). Peak 
inlet mass concentrations followed this 
springtime hydrograph peak: the maximum 
iron influx of 23.6 ppm occurred on May 1 
(fig. 3a), and the maximum manganese 



influx of 53.9 occurred on June 1 (fig. 
3c) . This sequence — peak mass 
concentrations following after peak flow 
volumes — could result from the springtime 
"flushing" of metals from the mine spoil. 

The wetland responded to the 
increased flow volumes by decreasing the 
percentage of iron removed from its inlet 
water. During the springtime increase of 
the inlet hydrograph (March 18 to April 
15) , the amount of mass removed dropped 
from approximately 10 kg (80-85% of the 
inlet mass) for each of the March 18 and 
April 7 periods, to 6.2 kg (43% of the 
inlet mass) for the April 15 period (fig. 



320 



B. 



c 

V 

o 

c 
o 
o 

JU 

c 



60 

50- 

40- 

30- 

20- 

10 



Fe 



■ ^-a^/"*-* 




r 25 
20 
15 5 

io £ 

5 



.-' to' <o' •-" eo' eg 1 m' d ot o' ol co' N' «" ,-' «>' ,-" <o' ro' 

.- CM CM OJ CM (M — — IT. ,- 

OND JFMAMJJ 



Sample Date 



c. 




Sample Date 



c 

v» 
o 

i_ 
-t-J 
C 
Q> 
O 

c 
o 
o 



3 

o 



c 
o 



c 

o 

c 
o 
o 



o 



60 

50 

40 

30 

20- 

10- 



Fe 




3M4 



/^-A-A-A- 



:^A 



25 
20 



15 1 
•10q 



•5 



i-A^ 



— I -J JT "«' '*Z?"JZ"2 — 7" ,J IT. 1 -J ^ ' ^. i ' — • — r . 

*- 00 <X> «- CO <N tf) ON O N O N tf) •- CO «- <D P) 
^- CN tN CN tN CM r- «- ,X «-- 

OND JFMAMJJ 



Sample Date 



D 

60 1 

50- 

40 

30 

20- 

10- 



Mn 




A-""~A-a' 



25 

20 



^ 15 i 



•10q 



- «o co -' co' N'n'd^oV co' n' n' „' «,' J »' k,' ' ° 

.- CM N CM N N ^ •■ ; " 

OND JFMAMJJ 

Sample Date 
A-±-A- A Concentration (ppm) 

Flow (gpm) 



Figure 3. — Inlet and outlet concentrations of iron (A & B) , and manganese (C & D) , 
documented for the 10-month study period. Plots of the inlet and outlet 
hydrographs are included for comparison of flow volumes with mass concentrations. 



2a) . The amount of iron removed during 
May 1 increased to 18 kg, but since inlet 
iron concentrations increased greatly 
during this period, this removal 
corresponded to only 55% of the total 
mass. 

The outflow hydrograph exhibited a 
lower, and more broad springtime peak, 
suggesting that the increased inlet 
volume did not immediately flow through 
the basin, but that a portion of the 
water was stored within the wetland, and 
then released at a more constant rate. 
Iron concentrations at the outlet 
increased with increasing outflow volume 
(March 18-May 1, fig. 3b) . 

During the April 15 period, when the 
inlet flow was at its highest and 
percentage iron removal at its lowest, 
manganese concentrations were slightly 
higher at the outlet than at the inlet. 
Mass budget calculations show an extra 
0.9 kg of manganese leaving the wetland 
during the April 15 period, and an extra 
1.3 kg leaving over the May 1 period. 
From May 16 through the end of the study 



period manganese was removed from the 
wetland flow. In fact, the greatest 
amount of removal (14.2 kg) coincided 
with the largest manganese influx (54.3 
kg) during the June 1 period (fig. 2b) . 

Outlet manganese concentrations were 
greatest during the intervals of June 1 
and June 16, reflecting the high inlet 
concentrations of those same periods 
(fig. 3d) . The peak outlet concentration 
of 43.6 ppm did not occur during the same 
interval as the peak inlet concentration 
however, instead the outlet peak appeared 
during following, June 16, period. 

The period from March 18 through 
April 15 was not the only period of 
rising inlet flow volumes: a small flow 
increase also occurred from November 21 
to December 8 (fig. 3c) . The response of 
the wetland to this flow increase was 
similar to its response during the 
springtime flow increase — iron removal 
dropped from 78% of the inlet mass during 
November 21 to 43% during December 8, and 
an extra 5 kg of manganese left the 
wetland over the December 8 period. 



321 



The effects of length and width on 
the removal of iron and manganese from 
the surface water were examined with an 
analysis of variance. Removal was 
calculated by subtracting the 
concentration at each grid point from the 
inlet concentration. Time effects were 
not examined with these tests; instead, 
all nineteen samples obtained from a 
particular location were considered to 
'replicate 1 the average chemical 
conditions at that area. While width was 
not found to be significant in the 
removal of either metal, length was 
significant in both cases, with iron 
being more affected. Duncan's multiple 
range test was used to examine the 
equality of the column means averaged 
over the three rows. It indicated that 
while removal did increase with length, 
the means for columns 7-11 (refer to fig. 
1 for sample grid design) were not 
significantly different from one another, 
similarly, the means for columns 5 and 6 
were also not significantly different. 

A second-order regression analysis 
of iron removal versus length produced a 
predicted curve (fig. 4), with a 0.962 
r-squared value. While a first-order 
relationship would also describe the data 
points, the quadratic curve provides a 
better fit and also suggests that the 
rate of iron removal slows during the 
time it takes water to flow from inlet to 
outlet. This is a likely possibility 
since reaction rates can slow as the 
reactants become more dilute. 



The 
removal o 
defined a 
to be a s 
test reve 
occurred 
10-11 ave 
as column 
length mu 
removal . 



^ 12 

E 

CL 10- 
CL 



importance of length in the 
f manganese cannot be as clearly 
s with iron. Length was found 
ignificant factor, but Duncan's 
aled that the greatest removal 
in columns 7-9, and that columns 
raged the same amount of removal 
s 5-6. Factors in addition to 
st be influencing manganese 
The first-order regression of 



(D 



8- 



O 

0) 4- 

c 

O 



y = -0.0008x + 0.206x + -3.758 
r 2 = 0.962 



20 40 60 80 100 120 

Length from Inlet (ft) 



manganese removal versus length (fig. 5) 
demonstrates this with an r-squared of 
0.53; length cannot explain all of the 
variability in manganese removal. A 
second-order fit did not significantly 
improve the r-squared value, and 
therefore is not presented here. 

The two-way interaction of length x 
width was significant as well, and a plot 
of the row means per column indicated 
that three surface locations (designated 
as [1,7], [1,8], and [1,9] in fig. 1), 
contained water with less iron and 
manganese, 37% and 20%, respectively, 
than water at the outlet. Samples from 
these areas were also distinguished by an 
enrichment in K, P, Na, Mg, and Ca. 
Similar water chemistry, i.e., high 
nutrient and low dissolved metal 
concentrations, was also found in the 
interstitial waters, especially from 
samples from the lower piezometers 
located immediately above the limestone 
layer. Based upon this information, two 
hypotheses can be made: 1) the high 
calcium and magnesium values at these 
three locations were due to the 
dissolution of limestone, which could 
have been mixed with the humic layer in 
these areas during wetland construction, 
and/or 2) interstitial water, rather 
than surface water might have been 
collected at these areas. The second 
possibility is likely since the wetland 
surface surrounding these areas can be 
described physically as a hummocky area 
where surface water was found in shallow 
depressions (0.5 to 1.0 inches deep) 
separated by emergent clumps of cattails, 
grasses, and saturated soil. Water 
collected as a surface sample might have 
actually flowed from the substrate at the 
time of collection. 

SUMMARY AND CONCLUSIONS 

An important question concerning 
this study is, of course: did the wetland 

Q.12 
Q. 



■o 

> 

O 

E 

<D 
(Z 

<D 
(0 
Q) 

C 
D 

en 

c 
o 



iz- 






10- 






8- 






6- 




y = 0.032x + -0.258 • 


4- 




r 2 = 0.530 m fl»^, — - 


2- 

0- 


• 

r- 


. . ._.. ., — — . _ — , 



20 40 60 80 100 120 

Length from Inlet (ft) 



Figure 4. --Plot of the relationship 
between iron removal and basin 
length. Each point represents 
the column average for the 10- 
month period. 



Figure 5. --Plot of the relationship 
between manganese removal and 
basin length. Each point 
represents the column average 
for the 10-month period. 



322 



effectively treat flow from the AMD seep? 
Effectiveness implies that effluent water 
quality meets standards set by the 
Surface Mining Control and Reclamation 
Act of 1977. Iron concentrations can be 
no greater than 7 ppm daily, or an 
average of 3.5 ppm monthly; manganese 
limits have been set to 4 ppm daily, and 
2 ppm for a monthly average (USEPA 1983) . 
Effluent from the wetland met the daily 
iron requirement at every sampling except 
for April 7, 15, and May 1, when the 
inlet flow volumes were higher than 
average. The monthly standard was met in 
every month but February, April, and May. 
Manganese concentrations remained above 
limitations throughout the entire period. 
In February, however, the mining company 
discontinued its downstream chemical 
treatment of the wetland effluent. 
Nevertheless, outflow from the polishing 
pond (fig. 1) still continued to meet 
iron and manganese requirements for the 
remainder of the study. 

Flow appeared to be an important 
variable affecting iron and manganese 
removal: increased flow volumes 
coincided with drops in in percentage 
removal of iron and increased outlet 
manganese concentrations. In addition, 
peak inlet mass concentrations, and 
therefore peak outlet concentrations, 
followed the peak springtime inflow. 

The importance of length as a factor 
which influences wetland treatment 
abilities agrees with findings from other 
researchers. A study by Wile et al., 
(1985) of a municipal waste wetland 
showed that one system with a length-to- 
width ratio of 75:1 constantly 
outperformed another with a 4.5:1 ratio. 
WAPORA (198 3) recommended a ratio of 
20:1. Within the present study, the 
increase of iron effluent concentrations 
during the spring (fig. 3b) suggests that 
the system was underdesigned for the 
maximum mass influx. The relationship 
between length and iron removal (fig. 4) , 
suggests that a lengthwise extension of 
the basin would increase iron removal. 
Although length was not as important in 
removing manganese, figure 5 suggests 
that a longer basin would also improve 
manganese removal. 

While outlet concentrations did not 
always show improvement in water quality, 
a few surface locations, and the 
interstitial waters, did exhibit 
consistently lower iron and manganese 
concentrations. Interaction with soil, 
and/or the underlying limestone layer, 
appears to be a common factor between 
these spots. A wetland designed with a 
more permeable substrate might provide 
greater contact between soil and water, 
provided water remains shallow. 

ACKNOWLEDGMENTS 

The authors would like to thank 
Horizon Coal Company, Beach City, OH, for 
its assistance and for the construction 



of the wetland. Tim Demko of P&N Coal 
Company, Inc. , of Punxsutawney, PA, also 
deserves thanks for his interest and 
suggestions. Partial funding for this 
study was provided by a fellowship from 
the Graduate and Professional 
Opportunities Program (GPOP) , and the 
Geology Department at Kent State 
University. 

LITERATURE CITED 

Brakersiek, D.L., H.B. Osborn, and W.J. 
Rawls. 1979. Field Manual for 
Research in Agricultural Hydrology. 
US Department of Agriculture, 
Agriculture Handbook 224. 550pp. 

Girts, M.A., R.L.P. Kleinmann, and P.M. 

Erickson. 1987. Performance data on 
Typha and Sphagnum wetlands 
constructed to treat coal mine 
drainage. Paper presented at the 
8th Annual Surface Mine Drainage 
Task Force Symposium, Morgantown, 
WV, April 7-8. 9pp. 

Girts, M.A., and R.L.P. Kleinmann. 1986. 
Constructing wetlands for treatment 
of mine water. Paper presented at 
the Society of Mining Engineers Fall 
Meeting, St. Louis, MO, September 7- 
10. 23pp. 



Girts, M.A. , and R.L.P. Kleinmann. 1986. 
Constructed wetlands for treatment 
of acid mine drainage: a 
preliminary review. p. 165-171 In 
The 1986 National Symposium on 
Mining, Hydrology, Sedimentology, 
and Reclamation, University of 
Kentucky, December 8-11. 

Hach. 1986. Water Analysis Handbook for 
the DR/3 Spectrophotometer. Hach 
Chemical Company. 4 3 6pp. 

Huntsman, B.E., J.G. Solch, and M.D. 
Porter. 1978. Utilization of a 
Sphagnum species dominated bog for 
coal acid mine drainage abatement, 
Geological Society of America (91st 
Annual Meeting) Abstracts, Toronto, 
Ontario, Canada, p. 322. 

Kadlec, R.H. . 1987. The use of peatlands 
for wastewater treatment. Session G 
In Session Abstracts, Symposium 
•87: Wetlands/Peatlands, Edmonton, 
Alberta, Canada, August 2 3-23. 
3 4pp. 

Kleinmann, R.L.P., R. Brooks, B. 

Huntsman, and B. Pesavento. 1986. 
Course notes on constructing 
wetlands for the treatment of mine 
water. PA Bureau of Mines. 2 9pp. 

LaBaugh, J.W., T.C. Winter. 1984. The 
impact of uncertainties in 
hydrologic measurement on phosphorus 
budgets and empirical models for two 
Colorado reservoirs. Limnol. 
Oceanogr. vol. 29. no. 2. p. 322- 
339. 



323 



USEPA. 1983. Neutralization of Acid Mine 
Drainage. EPA/ 600/2-83-001. 232pp. 

WAPORA, Inc.. 1983. Constructed 

wetlands. p. 3-114 to 3-121 In 
Bach, S.P., and J. R. Pilling 
(project officers) . The Effects of 
Wastewater Treatment Facilities on 
Wetlands in the Midwest. USEPA 
Technical Report 905/3-85-002. 

Wieder, R.K., and G.E. Lang. 1982. 

Modification of acid mine drainage 
in a freshwater wetland. p. 43-53 
In Symposium on wetlands of the 
unglaciated Appalachian region, West 
Virginia University, Morgantown, WV, 
May 26-28. 

Wieder, R.K., G.E. Lang, and A. E. 

Whitehouse. 1985. Metal removal in 
Sphagnum -dominated wetlands: 
Experience with a man-made wetland 
system. p. 353-364 In Proceedings 
of the Wetlands and Water Management 
on Mined Lands Conference, 
Pennsylvania State University, State 
College, PA, October 23-24. 

Wile, I., G. Palmateer, and G. Miller. 
1985. Use of artificial wetlands 
for wastewater treatment, p. 241- 
254 In Godfrey, P., E. Kaynor, S. 
Pelczarski, J. Benforado (eds.). 
Ecological Considerations in 
Wetlands Treatment of Municipal 
Wastewater. Proceedings of a 
Workshop, June 23-25 1982, 
University of Massachusetts, 
Amherst, MA. 



324 



CONSTRUCTED WETLANDS FOR ACID DRAINAGE CONTROL IN THE TENNESSEE VALLEY 



Gregory A. Brodie. Donald A. Hammer, and David A. Toml janovich 1 



Abstract--Constructed wetlands are often a 
preferred alternative to conventional methods of 
treating acid drainage at mine sites, coal 
preparation facilities, and coal-fired power 
plants. The Tennessee Valley Authority (TVA) has 
designed and constructed wetlands in Alabama and 
Tennessee to treat acid discharges from these 
sources. Between June 1985 and August 1987, seven 
wetlands were constructed to treat acid drainage 
at an inactive coal preparation plant and adjacent 
mined area and four wetlands were constructed at 
TVA coal-fired power plants. Although site- 
specific characteristics often restricted use of 
standardized methods, generic design, construc- 
tion, and operation guidelines have been 
developed. Treatment efficiencies have ranged 
from 82% - 99% removal for total iron and 9%- 98% 
removal for total manganese. Preliminary design 
guidelines for reguired treatment area were: ph < 
5.5 s.u., 2.0 m 2 /rag Fe, 7.0 m 2 /mg Mn; pH > 5.5 
s.u., 0.75 m 2 /mg Fe, 2.0 m 2 /mg Mn. Wetlands 
systems costs from design to operation averaged 
$12.18/m 2 of treatment area. 



INTRODUCTION 

Acid mine drainage is generated by 
the oxidation of iron sulfides in mine 
spoil, producing water that is acidic, 
with high concentrations of iron (Fe), 
sulfate, and other objectionable con- 
stituents, including manganese (Mn) , 
aluminum, suspended and dissolved 



L Listed alphabetically. Gregory 
A. Brodie is Environmental Engineer. 
Tennessee Valley Authority. Division of 
Fossil and Hydro Power. Chattanooga. TN; 
Donald A. Hammer is Senior Wetlands 
Ecologist. Tennessee Valley Authority . 
Division of Air and Water Resources. 
Norris. TN; David A. Tomljanovich is 
Biologist, Tennessee Valley Authority. 
Division of Air and Water Resources, 
Knoxville, TN. 



solids, color, and hardness (EPA 1971; 
Caruccio and Giedel 1985). Although 
stoichiometry may be somewhat different, 
acid drainage from ash ponds at coal- 
fired power plants is similar to acid 
mine drainage. 

Though effective, conventional acid 
drainage control techniques, such as 
chemical treatment or land reforming, 
are costly and generally require long- 
term maintenance. Constructed wetlands 
appear to offer an inexpensive, self- 
maintaining, long-term solution to 
treating acid drainage of moderate flows 
and chemical concentrations. The ability 
of wetlands to remove pollutants from 
acid drainage has been demonstrated by 
several investigators (Wieder et al. 
1984; Guertin et al. 1985; Brodie et al. 
1987). Various researchers have 
explored design and operational 



325 



parameters of constructed wetlands for 
treating acid drainage (Hammer and 
Kadlec 1983; Pesavento 1984; Huntsman et 
al. 1985; Girts and Kleinman 1986; 
Wieder and Lang 1986). 

The Tennessee Valley Authority's 
(TVA) fir6t constructed wetland for acid 
drainage treatment, known as Impoundment 
1, was located at a reclaimed coal 
preparation plant site in northeast 
Alabama (Brodie et al. 1987). Success 
of this wetland at meeting permit 
effluent limitations has led to the 
construction of 11 wetland treatment 
systems at TVA coal facilities and 
coal-fired power plants, and to an 
extensive research program on con- 
structed wetlands. This paper sum- 
marizes TVA's constructed wetlands 
treatment systems and offers preliminary 
design guidelines based on results to 
date. 



TECHNIQUES 

Site characteristics often restricted 
use of standardized methods; therefore, 
wetlands were designed for specific 
conditions. A generic description of 
pre-design investigations through 
wetlands operation follows. 



Pre-design 

State regulator 
onsite with TVA off 
alternative sites a 
Wetlands systems we 
as all effluent dis 
were to be met (i.e 
mg/L, total Mn < 2 . 
9.0 s .u. , and nonf i 
< 35.0 mg/L) . Ther 
ocassionally includ 
provide for chemica 
necessary. 



s generally met 
icials to review 
nd treatment options, 
re approved as long 
charge limitations 
. . total Fe < 3.0 

mg/L. pH = 6.0 - 
lterable residue 
efore system designs 
ed a final cell to 

1 treatment if 



Wastewater characterization and site 
hydrology were the two most important 
pre-design data needs. Pre-construction 
water quality sampling was conducted for 
all flows to be treated and any streams 
that would be affected by the wetlands. 
Analyses included pH, Eh, total and 
dissolved Fe and Mn. nonf ilterable 
residues (NFR) and total dissolved 
60lids, sulfate, aluminum, dissolved 
oxygen, manganese, zinc, selenium, 
mercury, and cadmium. Flow monitoring 
was incorporated into baseline site 
monitoring and compared to existing flow 
data and hydrologic modeling. Baseline 
population estimates of aquatic biota in 
receiving streams were estimated to 
provide a means of documenting stream 
recovery. 

Often it was necessary to locate the 
wetland off the immediate area or 
leased/owned surface, thus negotiations 
with landowners for initial access and 
long-term future control were conducted. 



TVA owned or purchased many of the 
wetland sites but also pursued other 
surface control arrangements including 
long-term leases and permanent easements. 

Regulations neither specifically 
addressed nor gave guidance on 
constructing treatment wetlands. 
Permits and approvals for wetlands 
construction required from State and 
Federal regulators included a National 
Environmental Policy Act review, 
floodplains (Executive Order No. 11988) 
review. National Pollutant Discharge 
Elimination System permit, surface 
mining permit, and a water engineering 
report. 

Topography was determined in 
sufficient detail to plan the number and 
location of cells to minimize cut and 
fill requirements dictated by a 
particular gradient. Because site 
regrading was usually an early step in 
wetlands construction, topography was 
completely altered, and detailed (e.g.. 
2-foot contour interval) topographic 
surveys were not warranted. 

Geology was evaluated to determine 
if the site overlay shallow bedrock or 
lacked suitable growth media. If 
necessary, sources of borrow and 
adequate growth media were identified. 
Flow patterns and depth of groundwater 
were determined to identify inflows or 
outflows that could affect water quality 
or hydrologic balance. 

Preferred sources of emergent 
vegetation for transplantation were 
nearby natural wetlands developed in 
similar quality water to avoid stress 
from abrupt changes in edaphic condi- 
tions. Cattail ( Typha ) , followed by 
Scirpus . Eleocharis . and Carex were the 
most tolerant, readily available species 
for transplantation. A rush Juncus was 
used with less success. Preliminary 
research results (S. R. Copeland, 
unpubl. data) suggested that Typha 
latif olia and Eleocharis quadrangulata 
provided higher radial oxygen loss than 
other common species, thereby enhancing 
substrate redox conditions to bind 
insoluble forms of metal precipitates in 
the substrate. 



Design and Construction 

Size of wetlands, . number of cells, 
spillways, and dike specifications were 
designed for 10-year. 24-hour storm 
event estimated from site flow monitor- 
ing or various methodologies (Lyle 
1987). Erosion and sedimentation 
control structure design and construc- 
tion (EPA. 1976; USDA. 1982). along with 
best engineering estimates and practices 
were used for those components having no 
design guidelines. 



326 



TVA's constructed wetlands ranged in 
size from 3.5 m 2 (38 ft 2 ) to 113.0 
m 2 (1216 ft 2 ) per average flowing 
liter per minute, and 2.0 m 2 (22 ft 2 ) 
to 41.0 m 2 (441 ft 2 ) per maximum 
flowing liter per minute. Design size6 
were dependent on water quality charac- 
teristics, storm flow hydrology, and 
land availability. Wetlands were 
designed to accommodate stormflow, then 
increased in size if very poor quality 
water was to be treated. Increasing the 
size of a wetland up to twice the 
stormflow design area only modestly 
increased costs and provided adequate 
treatment area. 

Wetlands shapes varied because of 
existing topography, geology, or land 
availability. Irregular shapes for 
wetlands cells enhanced natural appear- 
ance and provided hydraulic disconti- 
nuity. Configurations that increased 
velocities causing channelization, 
scouring, bank erosion, etc. were 
avoided . 

Number of cells for constructed 
wetlands was determined by site 
topography and hydrology. Level sites 
were amenable to large cells hydrau- 
lically chambered with rock or earthen 
finger dikes, large logs, vegetated 
hummocks, or other baffles. Steeper 
gradients required more grading or a 
system of several cells terraced down 
slope. 

Water depth and bottom slope were 
dependent on plant species, pollutant 
concentrations, freeze potential, and 
desired longevity of the system. Typha 
latif olia has been the preferred species 
in TVA constructed wetlands. Other 
plants used in shallow water include 
Scirpus . Juncus . Carex . Eleocharis . and 
Equisetum . Excessive water depths not 
only inhibited desirable species 
development, but promoted anoxic, 
reducing conditions in the water column 
seriously affecting the oxidation of 
iron and manganese. Shallow water 
depths which are subject to freezing in 
more northern climates are suitable in 
the Tennessee Valley. The primary 
advantages of shallow water is to 
enhance oxygenation and increase plant 
production. Potential disadvantages are 
reduced storage capacity and retention 
time. Average water depth in TVA's 
wetlands ranged from 15-30 cm (6-12 in) 
with some shallower and some deeper 
areas to provide for species 
diversification. A few deep pockets of 
one meter or greater were included in 
many cells to provide recharge zones and 
aquatic fauna refuge in drought events. 

Most wetlands were completed in 
early summer, although successful 
installations were completed as late as 
October. Wetlands construction began 
with clearing the site, followed by 
grading and dike construction, and 
importing suitable materials as 



necessary to meet design 
specifications. Brush was burned or 
pushed along the site perimeter to 
provide wildlife habitat. Spillways 
were either rock-lined or covered with 
non-biodegradable erosion control 
matting and planted with Scirpus . Carex . 
or grasses. Water level control or flow 
monitoring devices were incorporated 
into the spillway design and 
construction. 

Vegetation was hand-dug to obtain 
complete root balls/rhizomes. 
Transplantation was completed on the 
same day as digging, and plants were not 
subjected to extreme temperatures, 
drying, or wind during transport. Typha 
was set into the substrate at about nine 
plants /m 2 and stems broken over above 
the water level to prevent windfall and 
to stimulate new growth from the 
rhizomes. Bulrush clumps were simply 
placed in the desired location. 

Wetlands were fertilized with a 
phosphorous-potassium fertilizer such as 
0-12-12 at 400 kg/ha (353 lb/ac). 
Mosquito fish ( Gambusia af finis ) were 
stocked for insect pest control. 



Operation and Maintenance 

Post-construction activities 
included effluent monitoring, 
fertilization, and maintenance of dikes, 
spillways or other control structures 
and pest control. Effluent monitoring 
was performed several meters downstream 
of the final spillway so that any 
leakage was included in the sample. 
Monitoring requirements included pH, 
total Fe and Mn, and NFR. Additional 
water chemistry and biological 
monitoring were used to quantify 
wetlands treatment efficiencies and 
wetlands habitat benefits. After the 
first year, fertilization was done only 
if vegetation showed signs of nutrient 
depletion. Dike repair due to muskrate 
burrowing was required at one wetland. 
Army worm ( Simyra henrici ) infestations 
at another required control measures to 
prevent eradication of cattail. 



RESULTS 

A summary of characteristics and 
water quality parameters for TVA's 11 
constructed wetlands is presented in 
Table 1. Dates of initiated operation 
are based on the first time effluent 
monitoring began (i.e., usually within 
one week of initial dischange). Areas 
are given as surveyed inundated area. 
Influent water chemistry, in most cases, 
is based on at least one year of 
seasonal sampling of contributing 
seeps. Effluent monitoring, including 
flow, is generally based on twice 
monthly discharge permit sampling 
results from the date on which the 
wetland system began operating. 



327 



TABLE I 



TVA ACID DRAINAGE WETLANDS TREATMENT SUMMARY 



Wetlands 


Date 
Initiated 
Operation 


Area Number 
m 2 Cells pH 


Influent Water 
Parameters (mg/L) 
Fe Mn NFR 


pH 


Effluent Water 
Parameters (mg/L or L 

Fe Mn NFR 


/min) 
Fl 


ow 


Treatment 
m^/mg/m 
Fe 


area 
n 


System 


Ave 


Max 


Mn 


W C 018 


6-86 


4800 3 


5.6 


150.0 


6.8 




3.9 


6.4 


6.2 




70 


1495 


0.2 


4.2 


King 006 


10-87 


9300 3 


4.2 


153.0 


4.9 


40.0 










579 


2271 


0.2 


5.0 


Imp 4 


11-85 


2000 3 


4.9 


135.0 


24.0 


42.0 


4.6 


3.0 


4.0 


6.0 


42 


49 


0.4 


2.0 


950 NE 


9-87 


2500 2 


6.0 


II. 


9.0 


19.0 


6.6 


0.5 


0.2 


49. 1 


548 


1675 


0.7 


0.8 


R T - 2 


9-87 


7300 3 


5.7 


45.2 


15.4 




6.7 


0.8 


0.2 


2.0 


258 


681 


0.7 


2.5 


Imp 2 


6-86 


1 1000 5 


3.1 


40.0 


13.0 


9.0 2 


3.1 


3.4 


14.0 


0.8 2 


400 


2200 


0.7 


2.1 


Imp 3 


10-86 


1200 3 


6.3 


15.0 


5.0 


28.0 


6.8 


0.8 


1.9 


4.7 


87 


579 


I.I 


2.8 


W C 019 


6-86 


25000 3 


5.6 


17.9 


6.9 




4.3 


3.5 


5.9 




492 


6560 


2.8 


7.4 


950-1 8 2 


1976 


3400 3 


5.7 


12.0 


8.0 


20.0 


6.5 


I.I 


1.6 


5.4 


85 


541 


5.4 


5.1 


Imp 1 


5-85 


5700 4 


6.3 


30.0 


9.1 


57. 5 


6.5 


0.9 


2.1 


2.8 


55 


227 


5.6 


11.8 


Col 013 


10-87 


9200 5 


5.7 


0.7 


5.3 




6.7 


0.7 


15.5 




288 


408 


45.6 


6.0 



1 - one effluent sample to date 

2 - one sample, July 1987 

5 - from preconstruction instream sample 



Impoundment 1 was TVA's first 
constructed wetland, treating acid 
seepage from a coal slurry pond dike at 
the reclaimed Fabius Coal Preparation 
Plant in Jackson County, Alabama. 
Dominant vegetation is Typha . with 
Scirpus . Leersia . Juncus . Eleocharis . 
Utricularia . and Sparqanium among a 
total of 41 species present two years 

after construction. Since construction. 
Impoundment 1 has produced compliance 
quality effluent. 

Impoundment 4, also at the Fabius 
plant site, was built to treat acid 
seepage emanating from process water 
recirculation ponds. These ponds (pH = 
3.5 s.u.) were reclaimed in 1986 and the 
inflow to Impoundment 4 has been 
limited. Dominant plants are Typha and 
Scirpus . The original planting of 
Impoundment 4 took place in November 
1986. Few plants survived in spring and 
the wetland was replanted the following 
July. A sodium hydroxide (NaOH) 
treatment system was installed to 
augment the wetland treatment because of 
the extremely low pH of the seepage. 

950-1 and 2 was a two-cell 
sedimentation basin receiving mine acid 
drainage from the reclaimed TVA Fabius 
950 coal mine in Jackson County. A 
Typha marsh has naturally developed in 



the upstream cell with expansion into 
the lower cell. Treatment with NaOH, 
required from 1976 to 1984, has been 
discontinued. The discharge was 
released from NPDES permit monitoring 
requirements in 1987. 

Impoundment 2 is a series of 
constructed wetlands intermediate in a 
138 ha (341 ac) drainage basin receiving 
acid drainage from non-TVA abandoned 
mine land and the coarse refuse disposal 
area at the Fabius plant site. Effluent 
from the wetlands is treated with sodium 
hydroxide and discharged. Vegetation in 
the wetlands is predominantly Typha . 

Widows Creek 018. located at TVA's 
Widows Creek Fossil Plant in Jackson 
County, receives seepage from an 
abandoned ash disposal area. The 
wetland was adjacent to a leaking coal 
pile runoff pond (pH = 2.8 s.u.) that 
has caused adverse effects on water 
quality and vegetative development. 
Dominant vegetation is Typha . An NaOH 
treatment system was installed at this 
wetland for pH increase. 



Widows Creek 019 also receives acid 
seepage from abandoned ash disposal 
areas. Widows Creek Fossil Plant's 
operational needs at this site have 
resulted in plans to flood the wetland 



328 



and install a facility to pump water to 
an existing treatment system. Effluent 
data may reflect additional seepage 
within the wetlands system and should be 
viewed with caution. 

Impoundment 3 is located at the TVA 
950 coal mine and receives acid mine 
drainage. It was constructed to replace 
a chemical treatment system which 
operated from 1976 to 1986 and has 
produced compliance quality effluents 
since construction. Dominant vegetation 
is Typha . 

Rocky Top 2 is located at TVA's 
reclaimed Fabius Rocky Top coal mine in 
Jackson County. Inflow is acid mine 
drainage and dominant vegetation is 
Typha . 

950 NE is located adjacent to the 
TVA 950 coal mine and receives acid mine 
drainage from about 32 ha (79 ac) of 
reclaimed area. Dominant vegetation is 
Typha ; only minor discharge has occurred 
since construction. 

Kingston 006 is located at TVA's 
Kingston Fossil Plant in Roane County. 
Tennessee. It was constructed to treat 
acid seepage and runoff from active ash 
disposal areas. Dominant vegetation is 
Typha . About 20 cm (8 in) of 
high-calcium, minus 16 mesh limestone 
covered with about 30 cm (12 in) of 
spent mushroom compost for vegetative 
substrate was included in the final cell 
of the wetland system (B. G. Pesavento. 
pers . comm. ) . 

Colbert 013 is located at TVA's 
Colbert Fossil Plant in Colbert County, 
Alabama. It receives acid drainage from 
an indefinite source near an active ash 
disposal area. This wetland is still 
under development and is dominated by 
Typha and Scirpus . 



Water quali 
occurred at all 
constructed wet 
produced dramat 
apparently miti 
stubborn pollut 
regulatory limi 
achieved, cost 
a reduction in 
further metals 
adjustment . 



ty improvement has 

of the operating 
lands. Five systems have 
ic results and have 
gated some of TVA's most 
ion problems. Where 
ts were not entirely 
savings were realized in 
chemicals needed for 
precipitation or pH 



Total costs for construction of nine 
wetlands treatment systems are shown in 
Table 2. Overall average cost was 
$12.18/m 2 ($1.13/ft 2 ) of treatment 
area. Costs for 9 wetlands are best 
exemplified by the Impoundment 3 
project, which cost $40,000. Total cost 
consisted of about 20 percent for design 
and project management. 35 percent for 
equipment and supplies, and 45 percent 
for labor. Because TVA was annually 
spending $12,000 to $15,000 for 
chemicals and $10,000 for pond 



maintenance that failed to maintain 
complying discharges, the wetland 
system, with an annual operation and 
maintenance cost of $1,000. has proven 
cost-beneficial within one year. 



CONCLUSIONS 

Constructed wetlands offer a 
preferred alternative to conventional 
methods of treating acid drainage from 
certain coal-related sources. TVA has 
constructed 11 wetlands for treating 
acid drainage and has developed 
guidelines for design, construction, and 
operation of the systems. Six of these 
constructed wetlands allowed TVA to 
discontinue chemical treatment for Mn 
and Fe removal and pH adjustment. Two 
systems were under development but were 
expected to produce compliance quality 
effluents by mid 1988. The remaining 
wetland systems, although not treating 
water to compliance levels, removed 
significant amounts of metals from the 
influent, reducing chemical treatment 
costs. As more wetlands are constructed 
for acid drainage treatment, and as 
research results become available, 
design criteria will no doubt be 
improved. 

However, numerous wetlands treatment 
systems are planned or under 
construction in the coal and utility 
industries. Our experience suggests the 
following preliminary general guidelines 
for Fe and Mn treatment area 
requirements for desired discharge 
levels of Fe = 3 mg/L or less and Mn = 2 
mg/L or less 

Fe: 2 m 2 /mg<pH 5.5>0.75 m 2 /mg 

(21 ft 2 /ppm<ph 5.5>8 ft 2 /mg) 

Mn: 7 m 2 /mg<pH 5.5>2 m 2 /mg 

(75 ft 2 /ppm<pH 5.5>21 ft 2 /ppm) 

For pH levels of less than 5.5 s.u., 
suggested treatment area for Fe would be 
2 m 2 /mg/min (21 ft 2 /ppm) and for Mn 
7 m 2 /mg/min (75 f t 2 /ppm) . For pH 
levels above 5.5 s.u., suggested 
treatment area for Fe would be 0.75 
m 2 /mg/min (8 ft 2 /ppm) and Mn 2 
m 2 /mg/min (21 f t 2 /ppm) . For 
example, a seep with 50 mg/L Fe. 15 mg/L 
Mn, pH 5 . 6 s.u., and an average flow of 
113 L/min (30 gal/rain) would require: 



For Fe. the rate factor is 
m 2 /mg/min. 



75 



Area of treatment = 

0.75 m 2 /rag/min) (113 L/min) (50 mg/L) 
= 4237.5 m 2 

For Mn. area = 

2 m 2 /mg/min) (113 L/min) (15 mg/L) 
= 3390 m 2 

and the wetlands treatment system area 
should approximate 4200 m 2 (45,200 
ft 2 i.e.. about 1 acre). 



329 









TABLE 


2 












WETLANDS SYSTEM CONSTRUCTION 


COSTS 






Wetlands System 


Area 
ha 


X 
Equip 


X 

Labor 


X 
Overh 


Total 
($000' s) 


$/m 2 


$/ft 2 


R T - 2 


0.7 


32.6 


53.4 


14.0 


26.4 


3.58 


0.33 


Imp 2 


1.1 


70.2 


12.3 


17.5 


57.0 


5.25 


0.49 


V C 018 & 019 


2.9 


- 


- 


- 


209.0 


6.98 


0.65 


Imp 1 


0.5 


19.0 


49.6 


31.4 


42.9 


7.46 


0.69 


950 NE 


0.3 


38.1 


46.0 


15.9 


35.2 


13.80 


1.28 


Imp 4 


0.2 


28.2 


26.8 


45.0 


28.0 


14.12 


1.32 


King 006 


0.9 


28.6 


43.3 


28.1 


131.7 


14.21 


1.32 


Imp 3 


0.1 


34.8 


43.5 


21.7 


40.2 


32.03 


2.98 



average cost = $12.18/m 2 = $1.13/ft 2 



These treatment area estimations do 
not include storm flow hydrology and 
size of constructed wetlands must be 
increased, if necessary, to accommodate 
storm events and prevent dike or 
spillway damage. 



Magnolia Publ. Inc 
1032 pp. 



Orlando, Fla. 



Obviously all of 
generalize numerous f 
temperature dependent 
hydraulic loading, re 
surface area for micr 
length, width, depth 
system, etc., for whi 
information. Incorpo 
factors and developme 
relating influent val 
and flow to desired e 
recommended treatment 
loading, and retentio 
and will be reported 



the above 
actors, i.e., 

rate constants, 
tention time, 
obial growth, 
and slope of 
ch we have limited 
ration of these 
nt of an expression 
ues for Fe, Mn, pH, 
ffluent values and 

area, hydraulic 
n time is underway 
later. 



Although our results are encouraging 
and we suspect that properly designed 
and constructed wetlands treatment 
systems will function for long time 
periods, better documentation is 
obviously needed. TVA has initiated 
research addressing basic design 
guestions. including optimum substrates 
and vegetation, hydraulic and 
contaminant loading rates, treatment 
system capacity and longevity, storm 
event and ground water monitoring, and 
limestone bed design (Brodie et al. 
1988) . 



LITERATURE CITED 

Brodie. G. A.. D. A. Hammer, and D. A. 
Toral janovich, 1987. Treatment of 
Acid Drainage from Coal Facilities 
with Manmade Wetlands, p. 903-912. 
in K. R. Reddy and W. H. Smith, 
eds., Aguatic Plants for Water 
Treatment and Resource Recovery, 



Brodie. G. A.. D. A. Hammer, and D. A. 
Toml janovich. 1988. An Evaluation of 
Substrate Types in Constructed 
Wetlands Acid Drainage Treatment 
systems, in Proc. 1988 Mine Drainage 
and Reclam. Conf . . April 17-22. 
1988, Pittsburgh. Pa. 

Caruccio. F. T.. and G. Geidel. 1985, 
The Occurrence and Prediction of 
Acid Mine Drainage from Coal Strip 
Mines and Some Potential Answers to 
the Problem. Study Guide for a Mini 
Course taught at the 1985 Natl. 
Symp. on Surface Mining. Hydrology. 
Sedimentology, and Reclamation, 
Lexington. Ky., 41 pp. 

Girts, M. A. and R. L. P. Kleinmann, 
1986, Constructing Wetlands for 
Treatment of Mine Water, Presented 
at the 1986 Society of Mining 
Engineers Fall Meeting. St. Louis. 
Mo. Sep 1986. (Available from 
Robert Kleinmann. U.S. D.O.I. Bur. of 
Mines, Pittsburgh Res. Cent., 
Pittsburgh, Pa. 15236.) 

Guertin. deForest, J. C. Emerick. and 
E. A. Howard. 1985. Passive Mine 
Drainage Treatment Systems: A 
Theoretical Assessment and 
Experimental Evaluation. Colo. 
School of Mines. Golden, Colo. , 
73 pp. 

Hammer. D. E. and R. H. Kadlec. 1983. 
Design Principles for Wetland 
Treatment Systems, EPA No. 
PB-83-188-722. 



330 



Huntsman. B. E.. R. L. 
T. O. Tierman. 198 
Geochemical Consid 
Maintaining Manmad 
Constructed for Ac 
Abatement, p. 375. 
D. E. Samuel, and 
Wetlands and Water 
Mined Lands. The P 
Univ. . University 
pp. 



P. Kleinmann. 
5. Hydrologic and 
erations in 
e Wetlands 
id Mine Drainage 

in R. P. Brooks. 
J. B. Hill, eds.. 

Management on 
ennsylvania State 
Park. Pa.. 393 



Lyle. E. S». Jr.. 1987. Surface Mine 

Reclamation Manual. Elsevier Science 
Publ. Co.. Inc.. New York. 268 pp. 

Pesavento, B. G.. 1984, Factors to be 
Considered when Constructing 
Wetlands for Utilization as Biomass 
Filters to Remove Minerals from 
Solution, p. 45-49. in J. E.'Burris. 
ed.. Treatment of Mine Drainage by 
Wetlands. The Pennsylvania Univ., 
University Park. Pa., 49 pp. 



Wieder, R. K. , G. E. Lang, and A. E. 
Whitehouse, 1984. The Use of 
Freshwater Wetlands to Treat Acid 
Mine Drainage, p. 14-18, in Burris, 
J. E. ed.. Treatment of Mine 
Drainage by Wetlands, the 
Pennsylvania Univ.. University Park. 
Pa. , 49 pp. 

U. S. Dept. of Agriculture. 1982. Ponds- 
Planning. Design. Construction, Soil 
Conserv. Serv. . Agric. Handb. No. 
590. 51 pp. 

U. S. EPA. 1971. Acid Mine Drainage 
Formation and Abatement. Water 
Pollut. Control Res. Series. 
DAST-42. 14010 FPR 04/71. 

U. S. EPA. 1976. Erosion and Sediment 
Control. EPA-625/3-76-006. Vol. 1. 
102 pp. Vol. 2. 137 pp. 



Wieder. R. K. and Lang, G. E., 1986. 
"Fe. Al, Mn. and S Chemistry of 
Sphagnum Peat in Four Wetlands with 
Different Metal and Sulfur Input," 
Water. Air and Soil Pollut. 29: 
309-320. 



331 



THE SIMCO //A WETLAND: BIOLOGICAL PATTERNS AND PERFORMANCE 



OF A WETLAND RECEIVING MINE DRAINAGE. I 



Lloyd R. Stark, Ronald L. Kolbash, Harold J. Webster, 

9 

S. Edward Stevens, Jr., Kim A. Dionis, and Earl R. Murphy *• 



Abstract . --In 1985 a 3,000-m , three-celled wetland was 
installed at the Peabody Coal-American Electric Power Simco #A 
deep coal mine site near Coshocton, OH. The wetland included 
a limestone/compost substrate planted with Typha latifolia L. 
(cattail). The deep mine seepage pH is near 6.0, total iron 
80-2A1 mg/L, and acidity 15-389 mg/L. The percent reduction 
in iron concentration in effluent water relative to influent 
water of the wetland significantly exceeded pre-wetland percent 
reduction, improving from 28% during the first winter after 
construction to 62% in the summer of 1987. Further site 
modifications in 1987 elevated the percent iron reduction of 
the wetland to 70-85%. Iron reduction corresponds to greater 
wetland area, higher cattail density and increased plant 
coverage. Prominent plants in the wetland besides cattail are 
Leersia oryzoides (L.) Sw. (cutgrass) and several algal species. 
Cattail and cutgrass coverage increased from 1986 to 1987. Roots 
of cattail accumulated up to 5 times the iron content of roots 
from a control site. 



INTRODUCTION 

In the last 10 years over 100 wetlands have 
been constructed in the coal-bearing regions of 
Ohio, Pennsylvania, Maryland, and West Virginia 
in attempts to treat acidic mine water 
(Kleinmann and Girts 1987). The alternative can 
be costly: chemical treatment in settling ponds 
may cost a mining company as much as 
$50,000/year . Usually acidic mine water exceeds 



Federal regulations for iron (3 mg/L), manganese 
(2 mg/L), pH (6-9), or a combination of these. 
Sulfates and acidity can also be high. Wetlands 
have shown promise in lowering levels of iron 
and manganese (Wieder et al. 1982; Brooks et al 
1985). Monitoring the performance and 
composition of some of these wetlands will 
return useful knowledge on their potential. 



WETLAND CONSTRUCTION AND MODIFICATIONS 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation Conference 
sponsored by the American Society for Surface 
Mining and Reclamation and the U.S. Department 
of the Interior (Bureau of Mines and Office of 
Surface Mining Reclamation and Enforcement), 
April 17-22, 1988, Pittsburgh, PA. 

^L. R. Stark is a Research Associate, 
Department of Biology, The Pennsylvania State 
University, University Park, PA; R. L. Kolbash 
is the Administrative Assistant to the Vice 
President, Governmental Affairs, American 
Electric Power, Lancaster, OH; H. J. Webster is 
an Assistant Professor of Biology, The 
Pennsylvania State University, DuBois Campus, 
DuBois, PA; S. E. Stevens, Jr. is a Professor of 
Molecular and Cell Biology, The Pennsylvania 
State University, University Park, PA; K. A. 
Dionis is a Research Technician, Biotechnology 
Institute, The Pennsylvania State University, 
University Park, PA; and E. R. Murphy is a 
Senior Environmental Field Specialist, Peabody 
Coal Company, Zanesville, OH. 



The Simco //A underground mine began 
operation in 1970 and ceased operation on 
October 20, 1978. The location of the portal 
was a pre-existing strip bench operation of 
1961. This mine operated 3 conventional 
sections mining the Kittanning coal seam. 
Run-of-mine coal was delivered to the Columbus & 
Southern Ohio Electric Conesville Generating 
Station. 

Upon abandonment the mine was left unsealed 
and the highwall above the portal was not 
reclaimed to approximate original contour. In 
May 1979 the Division of Mines and the Mine 
Safety and Health Administration approved plans 
for sealing the Simco //A underground mine. The 
drift openings were sealed with 2A in. concrete 
or concrete bulkhead constructed with 
pilasters. One pilaster was used for widths 
under 16 ft. and two pilasters were used for 
widths greater than 16 ft. The seals were built 
into the floor, ribs, and top, and the tops were 
supported to make the areas safe prior to 
construction. All three openings at the Mine //A 
portal were sealed using this procedure, and the 
pre-existing highwall was subsequently 
backfilled. 



332 



In 1980, water treatment began for 
discharge seepage that developed from the mine 
near the base of the backfill. Discharge 
specifications were met by treating with soda 
ash briquettes. During the next several years, 
the flow increased to approximately 120 gpm 
while the water quality remained poor during the 
period 1980 through 1985: pH, 5.7-6.3; total 
iron, 145-241 mg/L; total manganese, 3.1-6.0 
mg/L; and acidity, 140-389 mg/L. 

As the flow increased it became 
increasingly difficult to treat the discharge 
effectively with soda asd briquettes. In an 
attempt to improve the treatment system and 
lower treatment costs, Simco/Peabody decided to 
use caustic soda and install an aeration system 
to elevate pH and enhance the settling of iron 
and manganese. 

From March 1980 through June 1985 
Simco/Peabody studied several options to 
effectively seal the mine to eliminate the 
discharge. One such method proposed sealing the 
Simco #4 underground mine with fixed Flu Gas 
Desulfurization Sludge from the Columbus & 
Southern Ohio Electric Conesville Generating 
Station. However, both the Office of Surface 
Mining Reclamation and Enforcement and the Ohio 
Environmental Protection Agency raised numerous 
questions and informational requests on the use 
of scrubber sludge. Concern was expressed about 
the long-term liability of the sludge if it were 
to become hazardous by decree or demonstration 
and also over the sludge fixation agent. 

It was thought that the proposed innovative 
sealing scheme may elicit a response from the 
regulatory agencies relieving liability and 
setting up a cooperative research effort; 
however, this did not happen. Since the Simco #4 
underground mine was under Notice of Violation 
(NOV) for its discharge, Peabody Coal developed 
cost estimates for the double bulkhead seal with 
center plug (the more conventional approach). 

Complicating the water discharge problem 
was an NOV issued by the Ohio Department of 
Natural Resources (ODNR), Division of 
Reclamation. The NOV issued May 1983 required 
Simco/Peabody to remove all treatment equipment, 
ponds, and totally reclaim the area. Otherwise, 
final bond release could not be obtained and the 
permit could be revoked. Abatement extensions 
were granted by ODNR while Peabody did further 
development on sealing methods. While the plans 
to reactivate the entries and construct 
hydraulic seals within the mine were developed 
and final construction details and bid requests 
were prepared, it was decided that such sealing 
was too costly, time consuming, and would not 
provide any assurance of complete effectiveness. 

Therefore, in April 1985, Peabody Coal 
contacted a wetland consultant, B. Pesavento, to 
perform a site evaluation of the abandoned Simco 
#4 underground mine. His analysis indicated 
that a wetland could be constructed that would 
sufficiently improve the discharge, bringing it 
within applicable discharge limits without the 
use of the existing treatment equipment. A 
proposal based upon Pesavento' s recommendation 



was submitted to the ODNR in July 1985. After 
careful review the ODNR permitted the 
installation of the wetland as a possible 
alternative to hydraulic sealing of the mine. 

Site conditions at the mine were conducive 
to construction of a wetland treatment system. 
The abandoned strip pit west of the backfilled 
portal area through which the discharge (120 
gpm) passed provided more than adequate area and 
sufficiently low slopes for construction of a 
wetland system. The pit floor was graded and a 
layer of crushed limestone (6 inches thick) was 
applied in the initial stages of wetland 
construction. The crushed limestone was covered 
with an organic-rich, deep-rooting medium (18 
inches) in which lime was incorporated. Typha 
rhizomes were planted with a density of 
3-4/m 2 . 

The wetland system was constructed in 
segments (cells) to provide proper gradients to 
restrict flow velocities and to promote uniform 
water dispersal throughout the wetland. 
Segmenting the wetland provided a gradual 
introduction of the discharge water into the 
wetland, allowing the vegetation to establish 
prior to the introduction of the entire 
discharge. 

Construction of the wetland was started on 
October 14, 1985 and completed 20 November 

1985. Three wetland cells separated by mixing 
pools were established (fig. 1). The total 
wetland area is about 3,000 m . Water 
quality, vegetation, and aerobic microbial 
populations have been monitored since June 1986, 
and are reported herein. 

Periodically the wetland has been limed 
and/or fertilized with phosphate (table 1) in an 
attempt to increase iron removal and promote the 
growth of algae. Such applications have now 
stopped. Despite the liming, pH was unaffected, 
and the iron removal efficiency of the wetland 
did not change significantly. However, the iron 
removal efficiency of the ditch below the 
wetland was distinctly elevated in the summer of 

1986, coinciding with the largest application of 
lime. Phosphate applications were relatively 
small. 

During the second winter after installation 
(1986-7) the water below the wetland at station 
DN02 required chemical treatment to remove more 
iron. At the time the water level and flow rate 
was judged to be too high in Cells 1 and 2, and 
the use of diversion ditches was recommended to 
effectively cut flow rate in half and lower the 
water depth. In addition, a series of hay bale 
dikes was recommended to be installed as baffles 
to increase retention time and promote bacterial 
activity. These changes were completed on 
September 1, 1987 (fig. 2). By this time the 
wetland had improved enough in efficiency to 
discontinue chemical treatment. However, recent 
analyses show the wetland to be performing at 
its highest efficiency, with respect to iron 
removal, since installation. 



333 



SEEP s 



CELL 




DN01 



ES01 



ES02 



ES03 



Figure 1. --Diagram of Simco #4 Wetland showing 
water sampling points at right. Dots 
represent permanent quadrat stations. 
Water sampling point DN02, not shown, is 
approximately 100-m below pond outlet. 
Cell areas: Cell 1: 1,008 m 2 ; Cell 2: 
86A m 2 ; Cell 3: 1,200 m 2 . 



Table 1.- -Dates and amounts of agricultural 
lime and phosphate 

(di-ammonium phosphate: 0-45-0) applied 
to the wetland. 



Date 


Lime 


Fertilizer 




(lbs) 


(lbs) 


Dec 1985 


-- 


200 


Feb 1986 


-- 


200 


Mar 1986 


— 


200 


Jul 1986 


8,000 


— 


Nov 1986 


950 


200 


Dec 1986 


1,000 


200 


Jan 1987 


-- 


150 


Apr 1987 


-- 


150 


Jul 1987 


-- 


150 


Sep 1987 


— 


150 



METHODS 

Water Chemistry 

Water was collected by Peabody Coal at 2-3 
week intervals at the seep, the outfall of each 
of the three wetland cells, and 100 m downstream 
of the wetland, just prior to entering the 
treatment pond. Water was analyzed by the 
Peabody Coal Lab for acidity, total iron, total 
manganese, and sulfate. In the field, pH, 
temperature, and the water flow rate discharging 
from the wetland were taken. Acidity was 
determined using an Orion 407/A pH meter by 
titration (APHA 1980). Metal concentrations 
were determined with a Perkin/Elmer 4000 and 
5000 atomic absorption spectrophotometer using 
standard techniques. Sulfates were determined 
turbidimetrically using a Hach Turbidimeter 
Model 2100A (through July 1987) and a Hach 
Ration Turbidimeter Model 18900 (August 1987 
on) . 

On a quarterly basis The Pennsylvania State 
University (PSU) collected water samples at the 
permanent sampling stations established within 
the wetland (fig. 1). These were then filtered 
(0.45 um) and acidified (HN0 3 ) in the field 
and analyzed by atomic absorption 
spectrophotometry using a Buck 200 AA for 
determination of dissolved iron and manganese. 

Vegetation 

Vegetation was surveyed by PSU at each of 
the 40 permanent sampling stations on a 
quarterly basis beginning in July 1986. Canopy 
coverage was estimated (Daubenmire 1970) in the 
summers of 1986 and 1987, the first two growing 
seasons for the wetland. In 1987 algal presence 
(January) and cover (April) in each quadrat were 



334 



SEEP 



CELL 1 




DITCH 



CELL 2 



MIXING POOL 



HAY BALE DIKE 



POND 



Figure 2. --Diagram of Simco #4 Wetland after 

site modifications of August 31, 1987. Hay 
bale dikes were spaced at approximately 
35-ft. intervals. 



recorded. In addition, the number of cattail 
shoots was counted in the quadrats in April and 
August 1987. 

Metals in Typha Structures 

Whole cattail plants were collected in 
April and August 1987 and returned to the lab in 
plastic bags. These plants were then separated 
into leaf blades; leaf bases; inner, outer and 
entire rhizomes; and roots. Damaged or dead 



plant parts and visible contaminants were 
removed, followed by cleaning in tap water, with 
vigorous rinses and gentle brushing to remove 
dislodgeable surface contaminants. Samples were 
then oven-dried at 70°C, ground in a Wiley 
mill, and stored in sealed glass jars. Wet acid 
digestion was used, modified from the procedures 
in Allen et al. (1986). The dried plant 
material (500 mg) was acid digested at 160°C, 
first with 10 mL dH 2 (distilled water) and 5 
mL concentrated H2SO4 for 20 minutes, then 5 
mL concentrated HNO3 was added and digestion 
continued for 10 minutes. After cooling, 
samples were filtered (Whatman GFC glass 
filters), diluted to a volume of 100 mL using 
deionized distilled H2O, and stored at 2°C. 
Two aliquots of each digestate were analyzed by 
atomic absorption spectrophotometry for iron and 
manganese concentrations. Each value reported 
represents an average of the two aliquots. 
Control plants were collected from a 
non-impacted site near DuBois, PA and analyzed 
as above. 

Microbial Studies 

The microbiological work has essentially 
been an attempt to discover the major 
nutritional types of microorganisms present in 
the wetland, their total number, and their 
diversity. Until recently heterotrophic 
microorganisms were not thought to occur in high 
iron-loaded waters. However, Wichlacz (1980) 
showed that acidophilic heterotrophs were 
present in acid mine waters, a result confirmed 
by McHerron (1986). 

Samples were collected in September 1986 
and in January, April, and October 1987 at 20 
locations along the length of the wetland. 
Depending on the nature of the sediments, 
samples were obtained in one of two ways. If 
the sediments were solid enough to hold a shape, 
solid cores were taken with a 3-cm-diameter 
metal cylinder. If the sediments were loose and 

watery, samples were collected by holding a 
sterile test tube just below the surface of the 
sediment, in front of the flow. When the tube 
was uncapped, the sediments flowed into the tube 
with as little disturbance as possible. The 
samples were kept on ice during transportation 
to the laboratory. 

For solid core samples, the outer portion 
of the core was removed with a sterile razor 
blade to expose the inner, presumably 
undisturbed, portion. A subsample of 2.5 g was 
blended in a Waring blender with 1% sodium 
pyrophosphate (Balkwill 1977). The purpose of 
this step was to release the microorganisms from 
the soil particles; sodium pyrophosphate aids in 
breaking up the soil aggregates. If the test 
tube samples were watery, the sediments were 
blended with enough sodium pyrophosphate to give 
an approximate 1% solution. From this point on, 
both types of samples were treated in the same 
manner . 

The suspensions were immediately diluted 
using sterile distilled in a series of 1:00; 
1:1,000; 1:10,000; and 1:100,000. Each dilution 
tube was placed on a vortexing machine before 
any liquid was drawn from it so the particles 
did not have time to settle out. All transfers 



335 



were made with micropipettes. Each dilution was 
then plated out in triplicate on seven different 
kinds of agar media. The media used ranged from 
very simple and defined (agar water and 
inorganic media) to complex and undefined (PTYG 
and BHI), in order to select for a gradient of 
nutritional types (Balkwill and Ghiorse 1985). 
Both iron-supplemented and 

manganese-supplemented inorganic media were used 
to select for iron and manganese oxidizers. 
Plates were surface inoculated and incubated at 
27°C. 

Colonies were counted and described for 
plates on which 30-300 colonies developed. The 
counts included fungi, actinomycetes , and 
bacteria. The final counts and descriptions 
were made at 15-30 days, depending on the type 
of medium used. A series of counts was also 
made within this time period to estimate growth 
rates (data not presented). Colony types were 
distinquished on the basis of relative size, 
color, conformation, border, and elevation. The 
presence of iron or manganese oxidizers was 
determined visually by the formation of a deep 
blue color after flooding the colonies with 0.2% 
tetramethyl benzidine in 2M acetic acid (Stone 
1984). The iron and manganese oxidizers were 
isolated from these plates for further study. 
Gram stains were performed on the predominant 
colony types, and cellular morphology was 
described. 



Table 2. --Average pH values by season of each 
wetland cell and ditch outfall; N=5 for 
pre-wetland means, N=4-7 for post-wetland 
means . 



Location Pre-wetland Winter Spring Summer 
(1983-1985) 1985-6 1986 1986 



Seep 
Cell 1 
Cell 2 
Cell 3 
Ditch End 



6.16 



5.1 



6.45 


6.40 


6.40 


6.07 


6.52 


6.40 


5.92 


6.46 


6.22 


6.03 


6.40 


6.22 


6.07 


6.50 


6.24 



Fall Winter Spring Summer Fall, 



1986 



1986-7 1987 



1987 



Seep 6.46 6.34 

Cell 1 6.16 6.48 

Cell 2 6.16 6.48 

Cell 3 6.47 6.36 

Ditch End 6.16 6.50 



1987' 



6.25 


6.36 


6.39 


6.45 


6.33 


6.29 


6.40 


6.31 


6.27 


6.27 


6.19 


5.99 


6.43 


6.26 


6.18 



RESULTS 

Water Chemistry 

Flow . Flow rate measured at the outflow of 
the wetland ranged from 100-160 gpm from 
September 1986 to August 1987, but declined to 
50-65 gpm in September 1987. Lower than normal 
precipitation in 1987 in combination with the 
site modifications of late August 1987 may have 
contributed to this flow reduction. 

pH . Prior to wetland installation the pH 
on average decreased from 6.16 to 5.88 between 
the seep (DN01) and the final sample point 
(DN02; table 2). Following wetland installation 
the pH has remained relatively constant. 
Although too early to detect a pattern, 
following site modifications of August 1987 the 
pH of water leaving Cell 3 was the lowest (5.99) 
since wetland installation. 

Iron . Total iron at the seep ranged from 
80-241 mg/L from 1983-1987, but was usually 
100-120 mg/L (table 3). Pre- and post-wetland 
iron levels can be compared at the end of the 
100 m-long ditch below the wetland (station 
DN02, approximately 100-m below the pond outlet) 
which was sampled prior to and after wetland 
installation. "Efficiency" (or more properly 
the percent reduction in iron concentration of 
effluent water relative to influent water) in 
removing iron is defined as: 

Total Fe cone, at DN01 - Total Fe cone, at DN02 
Total Fe cone, at DN01 



after site modifications of August 31, 1987 



Winter = Dec-Feb 
Spring = Mar-May 
Summer = Jun-Aug 
Fall = Sep-Nov 



Prior to wetland installation the total iron 
concentration was reduced by an average of 32% 
between DN01 and DN02 (180-122/180). During the 
first winter following installation the iron 
reduction efficiency for the wetland (at DN02) 
was 35% (table 4). This efficiency rose through 
the spring and summer to 81% in July 1986, 
averaging 75% in its first summer and 70% in the 
second summer. The efficiency in the second 
winter (1986-7) exceeded that of the first 
winter, 53% to 35%, as did the efficiency of the 
second spring to the first, 62% to 47%. Using 
this efficiency index, this wetland is improving 
over time (fig. 3). 

Because station DN02 is 100-m below the 
wetland, a more accurate gauging point for 
wetland efficiency is ES03, which is at the 
outfall of the third (final) wetland cell. The 
efficiency at ES03 reflects the effect of the 
wetland system proper on water quality. The 
cumulative percent reduction at ES03 (table 4) 
climbed steadily from 28% to 56% in the summer 
of 1986 and remained near that level until the 
following summer, when it rose to 62%. Thus, 
(a) there was not a significant dropoff in the 
wetland' s efficiency in the reduction of iron 
over the winter, and (b) the level of iron 



336 



Table 3. --Average total iron (mg/L) by season of 
each wetland cell and ditch outfall; 
N=5 for pre-wetland means, N=4-7 for 
post-wetland means. 



Table 4. --Average percent reductions in total 
iron concentration by season of each 
wetland cell, the ditch, cumulative (cum.) 
for the wetland proper, and end of ditch; 
N=4-7 for each season. 



Location Pre-Wetland 
(1983-1985) 



Winter Spring Summer 
1985-6 1986 1986 



Component Winter Spring Summer Fall 
1985-6 1986 1986 1986 



Seep 180 

(range: 


144-241) 


116 


106 


127 


Cell 1 
Cell 2 




14 
6 


12 
12 


22 
20 


17 

14 


Cell 1 







99 


93 


98 


Cell 3 




12 


18 


27 


29 


Cell 2 







93 


81 


77 


Ditch 




10 


16 


42 


22 


Cell 3 







84 


65 


56 


Cum. wetland 


28 


37 


56 


53 


Ditch End 


122 




75 


55 


32 


Cum. , end 
of ditch 




35 


47 


75 


59 


















Fall 
1986 


Winter 
1986-7 


Spring 
1987 


Summer 
1987 


Fall 
1987* 
















Component 


Winter 
1986-7 


Spring 
1987 


Summer 
1987 


Fall 














1987^ 


Seep 


144 


113 


100 


87 


100 


Cell 1 




17 


21 


22 


38 


Cell 1 


123 


100 


78 


68 


62 


Cell 2 




15 


17 


17 


41 


Cell 2 


105 


86 


63 


59 


59 


Cell 3 




30 


33 


42 


56 


Cell 3 


74 


60 


42 


34 


26 


Ditch 




8 


13 


22 


26 


Ditch End 


58 


52 


38 


26 


19 


Cum. wetland 
Cum. , end 


51 
53 


55 
62 


62 
70 


73 














81 


after 


site modifications of August 31, 


1987 


of ditch 













Winter = Dec-Feb 
Spring = Mar-May 
Summer = Jun-Aug 
Fall = Sep-Nov 

reduction was much greater in the second winter 
(51%) than the first winter (28%). In the 
interval between the first and second winters 
this constructed wetland experienced its first 
growing season. 

A knowledge of the efficiency of each 
wetland cell is valuable in further evaluating 
the wetland system. The percent reduction of 
Cell 1 is the efficiency up to ES01; the percent 
reduction of Cell 2 uses ES02 and ES01 as 
reference points; and the percent reduction of 
Cell 3 uses ES03 and ES02 as reference points. 
Prior to the site modifications, Cells 1 and 2 
showed peak efficiencies in the summers of 1986 
and 1987. The variation here is high, but 
3-month summer averages for Cells 1 and 2 are 
22% and 20%, respectively (table 4). The 
efficiencies for these two cells decreased 
slightly after the summer (3-month averages of 
17% and 14%). Cell 3 efficiency, however, has 
increased steadily since installation, 
performing at twice the efficiency of Cell 2. 
The improvement of the wetland system proper 
from 28% (1986) to 51% (1987) in winter 
efficiency is in large part due to the high 



after site modifications of August 31, 1987 



Winter = Dec-Feb 
Spring = Mar-May 
Summer = Jun-Aug 
Fall = Sep-Nov 



winter efficiency of Cell 3. The decline in 
cumulative efficiency at point DN02 in the fall 
and winter of 1986-7 is therefore not due to the 
action of the wetland itself, but to the 
decreased removal in the ditch between ES03 and 
DN02. Iron removal by the ditch ranged from a 
high of 42% in summer 1986 to only 8% in winter 
1986-7 (table 4). 

Water samples were taken within the wetland 
at each of 40 stations on a quarterly basis. 
Representative results for August 1987 (fig. 4) 
indicate that no sharp transition zones of iron 
reduction exist; a gradual decrease in iron 
content in the water occurs from inlet to 
outlet. Higher iron levels are present on the 
right-hand (highwall) side of the wetland; this 
is probably due in large part to the much 
greater flow and channeling observed on this 
side prior to the August 31, 1987 modifications. 



337 




15SEP85 15N0V85 15JAN86 I5MAR86 15MAY86 15JUL86 15SEP86 15NOV86 15JAN87 15MAR87 15MAY87 15JUL87 15SEP87 15NOV87 15JAN88 

DATE 
STATION -a — a — a DNO 1 ^-.-cy-.o DN02 * — * — * ES03 



Figure 3. --Total iron concentrations at DN01 (seep), ES03 (wetland proper 
outlet), and DN02 (100-m below wetland). 



The introduction of hay bale dikes and the 
diversion of half the water around each of Cells 
1 and 2 (performed between the summer and fall 
of 1987) resulted in a further reduction of iron 
in the water. Although we cannot predict how 
the wetland would have performed without 
modification, we can compare patterns between 
the summer-fall of 1986 and the summer-fall of 
1987. In the fall of 1986 the cumulative 
percent reduction in the iron concentration 
after passing through the wetland was 53% (table 
4), 3% lower than the summer 1986 efficiency. 
However, the iron reduction efficiency in the 
fall of 1987 was 73%, 11% higher than that of 
the preceding summer. Each wetland cell 
increased in efficiency (a direct increase of 
14% to 24%; table 4), with the middle cell 
showing the most dramatic improvement. Prior to 
site modifications, summer had been the season 
of peak performance for iron reduction (56% and 
62%). However, in the fall of 1987 after 
modification, an all-time seasonal high of 73% 
was recorded (table 4). 

Manganese . Total manganese in the seep 
water ranged from 1.66-6.02 mg/L from 1983-1987, 
averaging 2.12 since 1985 (N=43). Wetland 
outlet water over the same period has ranged 



from 1.81-6.09 mg/L, averaging 2.28 mg/L 
(N=43). Thus, the Mn concentration in the water 
increased by 8% between seep and wetland outlet, 
with the elevation occurring in Cells 1 and 2. 
Although the manganese levels are a minor 
concern due to their low levels, the probable 
explanation of the slight elevation in Mn over 
the first two wetland cells was found to be a 
small surface side seep of <1 gpm alongside Cell 
1 that contained about 5 mg/L Mn and negligible 
Fe. 

Acidity . Pre-wetland acidity at the seep 
and the ditch outfall averaged 207 and 187 mg/L, 
respectively: an improvement of 10%. The effect 
of the wetland was to reduce acidity (table 5): 
from the summer of 1986 forward the wetland has 
lowered acidity by at least 63%. Cell 3 has 
been the most effective cell in reducing acidity 
in five of the seven seasons, followed by Cell 
1. 

Sulfate . No data on sulfates are available 
prior to wetland installation. Since late 1985, 
sulfates in the seep water have averaged 1 , 226 
mg/L (N = 40; range = 964-1,841 mg/L). The 
wetland has lowered this level by 7%, (N = 40; 
range = 425-1,931 mg/L). 



338 



Vegetation 

Following the initial Typha latifolia 
transplants the wetland has basically been 
allowed to establish on its own. Seventeen 
species of plants have appeared in the wetland, 
but only a few account for a significant biomass 
based on coverage (table 6): Typha latifolia , 
Leersia oryzoides , and mixed algal populations 
including primarily Oscillatoria sp., 
Chlamydomonas polypyrenoideum , and Euglena sp. 



SEEP 



CELL 1 




These three elements (i.e., the two higher 
plants and the algae) accounted for 87% of the 
total plant cover in 1986, and for 92% in 1987. 

Total cover in 1986 was 45%; in 1987 it 
increased to 83%, in large part due to increased 
growth of the three elements above. The 
cutgrass ( Leersia ) was a relatively minor 
element of the vegetation in 1986, but in 1987 
attained one-half the cover of the cattails. 
Cell 3 has the highest coverage of flowering 
plants, while Cell 2 cover increased 
dramatically in 1987. Algae were quite abundant 
in Cell 2, in part due to the frequent open 
spaces with reduced higher plant cover. The 
cattail cover in Cell 1 did not increase between 
1986 and 1987; muskrats were present in this 
cell, fed on cattail rhizomes, and were trapped 
out intermittently. 

Cattail shoot counts in August 1987 yielded 
the following estimates of shoot density per 
wetland cell (N is the number of 0.5-m 
quadrats sampled): Cell 1: 6.5lm^ (N=8); 
Cell 2: 
(N=9). 
Cell 3. 



12.0/m 2 (N=6); Cell 3: 18.7/m 2 
Clearly cattails are more numerous in 



Table 5. --Average total acidity (mg/L) by season 
of each wetland cell and ditch outfall; 
parenthetical values are the percentage 
reduction in acidity from the seep; 
N=5 for pre-wetland means, 
N=4-7 for post-wetland means. 



Location Pre-Wetland Winter Spring Summer 
(1983-1985) 1985-6 1986 1986 



Seep 


207 


95 


52 


92 


Cell 1 


--- 


82 


38 


57 


Cell 2 


- — 


76 


36 


32 


Cell 3 


--- 


70 


26 


15 


Ditch End 


187 
(10%) 


65 
(32%) 


22 
(58%) 


1 
(99 



Fall Winter Spring Summer Fall, 
1986 1986-7 1987 1987 1987* 



POND 



Seep 


88 


54 


57 


46 


61 


Cell 1 


64 


38 


25 


24 


3 


Cell 2 


59 


27 


22 


15 


17 


Cell 3 


41 


20 


14 


3 


-4 


Ditch End 


20 
(77%) 


20 
(63%) 


15 
(74%) 


6 
(87%) 


-3 
(100%) 



after site modifications of August 31, 1987 



Figure 4. --Dissolved iron (mg/L) in water 

samples taken at 40 evenly spaced stations 
within the wetland, August 1987. 



Winter = Dec-Feb 
Spring = Mar-May 
Summer = Jun-Aug 
Fall = Sep-Nov 



339 



Table 6. --Coverage of plant and algal species by cell in 1986 and 1987. A "--" 

denotes absent, and a "+" denotes present but with less than 1% mean cover. 



Species 



Cell 1 
1986 1987 



Cell 2 
1986 1987 



Cell 3 
1986 1987 



Overall 
1986 1987 



Typha 29 
latif olia 
(cattail) 



28 



20 48 



39 



51 



30 



42 



Leersia 

oryzoides 

(cutgrass) 

Lemna minor 
(duckweed) 



1 20 



35 



22 



Equisetum 

arvense 

(horsetail) 

Phalaris 1 

arundinacea 

(reed canary grass) 

Polygonum 
sagittatum 
(tear -thumb) 

Verbena 
hastata 
(vervain) 

Salix sp. 
(willow) 

Alisma sp. 
(water plaintain) 

Ailanthus 
altissima 
(tree of heaven) 

Epilobium sp.-- 
(willow-herb) 

Algae 9 



12 



24 



12 
~83~ 



TOTAL 



41 



55 



27 



98 



60 



97 



45 



•including: mixed populations of Oscillatoria sp.; 2 Euglena sp.; 
Chlamydomonas polypyrenoideum ; and Microspora pachyderma . 



In the winter, 22 of 40 quadrats had 
visible algal populations, with 12 of these in 
Cell 3. Overall algal cover in the spring was 
41%, with Cell 1 = 54%, Cell 2 = 30%, and Cell 
= 37%. The algae are a prominent vegetational 
element on the substrate and water surfaces 
year-round. 

Plant Tissue Metal Levels 

Cattail roots at the Simco #4 wetland had 
the highest iron concentrations, followed in 
sequence of decreasing concentrations by outer 
rhizomes, entire rhizomes, leaf bases, leaf 
blades, and inner rhizomes (table 7). Control 
plants had higher iron values in the outer 
rhizome samples than found in the experimental 



(i.e., Simco /M wetland) plants. Control root 
values were greater than three of six 
experimental values. Although no apparent 
surface mining has occurred in the immediate 
watershed where controls were taken, it is 
possible that some past disturbance has resulted 
in elevated iron levels in the substrate. 

All roots and outer rhizome samples were 
iron-red and had iron encrustations on the 
surface that were difficult to remove from the 
epidermis. Inner rhizome tissues and 
hand-sectioned root tissues were not 
discolored. These observations correlate with 
the higher iron values observed in this study. 
Taylor et al. (1984) demonstrated that iron 
coatings (plaques) form on the surface of 



340 



Table 7. --Average iron and manganese concentrations (mg/g dry weight) 
in plant structures of Typha latifolia from the Simco //4 wetland 
and a control site near DuBois, PA; numbers of plants sampled in 
parentheses. 



Structure 



Location 
(Cell No.) 



Fe 
Jan Apr 
'87 '87 



Mn 



Jan 
'87 



Apr 
'87 



Leaf Blade 



Leaf Base 
(sheath) 



Inner Rhizome 



1 -- 0.86(2) 

2 -- 1.02(2) 

3 — 1.27(3) 
Control 0.65(3) 

1 14.69(1) 0.78(2) 

2 15.50(1) 0.92(3) 

3 -- 1.20(2) 
Control 0.26(3) 

1 0.55(4) 0.84(4) 

2 0.29(3) 2.70(2) 

3 0.33(4) 1.52(5) 
Control 0.39(1) 



0.47(2) 
0.41(2) 
1.86(3) 
.07(2) 



0.02(1) 
0.38(1) 



0.23(2) 
0.13(3) 
0.52(2) 
0.92(3) 



0.12(4) 0.13(4) 

0.05(3) 0.07(2) 

0.21(4) 0.18(5) 
0.24 



Outer Rhizome 1 
2 
3 
Control 



8.34(3) 22.16(2) 
12.77(3) 7.95(3) 
11.52(2) 2.51(2) 
14.51(2) 



0.31(3) 0.18(2) 

0.05(3) 0.09(3) 

0.12(2) 0.27(2) 
0.41 



Entire Rhizome 



Root 



1 3.61(4) 7.71(4) 

2 6.03(4) 1.73(2) 

3 2.46(3) 6.24(2) 
Control 0.60(1) 

1 9.11(4) 27.82(3) 

2 21.95(3) 41.65(2) 

3 107.04(2) 12.63(1) 
Control 21.40(2) 



0.06(4) 0.13(4) 

0.03(4) 0.09(2) 

0.12(3) 0.15(2) 

0.48(1) 

0.11(4) 0.26(3) 

0.08(3) 0.10(1) 

0.20(2) 0.61(1) 

0.65(2) 



cattail roots when ferrous iron is oxided and is 
deposited on the roots and in the adjacent 
rhizosphere. Macfie and Crowder (1987) found 
that iron plaque formation in cattails was 
positively correlated with extractable iron in 
the substrate and pH. Consequently, higher iron 
values would be expected in surface tissues than 
in inner tissues. 

Roots and rhizomes have higher iron 
concentrations than foliar tissues (table 7). 
Other researchers have found the same pattern in 
cattails (Taylor and Crowder 1983; Adriano et 
al. 1984). In contrast, foliar samples of 
cattails had higher manganese concentrations 
than rhizome and root samples (.table 7). 
Adriano et al. (1984) found no seasonal patterns 
for manganese accumulation. However, their 
study demonstrated rhizomes had much lower 
concentrations of manganese than shoots. 

Microbial Studies 

Respectably high numbers of aerobic 
heterotrophs (the average viable cell count for 
the entire wetland was 1.4 x 10° colony 
forming units/mL sediment) occurred in the 
wetland. Typical trends in the numbers of 



aerobic heterotrophs through the wetland from 
the seep to the outfall are shown in fig. 5. An 
obvious bloom in the total population of aerobic 
heterotrophs (averaging about two orders of 
magnitude in increased population) was seen in 
the spring samples. Bacterial numbers were 
lowest in the winter samples. The highest 
viable cell counts were found on agar-water and 
brain heart infusion agar. To put these results 
into context over a 1-year period, the numbers 
of bacteria in the wetland samples were about 
two orders of magnitude lower than those found 
in rich agricultural soils, about the same as 
those found in eutrophic lakes, and three orders 
of magnitude higher than those found in 
oligotrophic lakes. However, species diversity 
of the aerobic heterotrophs was variable and 
relatively low over the wetland, averaging only 
between 9 and 25 distinct colony types. Also, 
iron- and manganese-oxidizing bacteria were 
quite variable (from none detected to over 
10°/mL of sample) through the wetland (data 
not shown). This was surprising because the 
decline in soluble iron from the seep to the 
outfall was relatively smooth, while the 
manganese content increased slightly. Black 
water with the characteristic smell of H2S and 
that of mercaptans was noted, especially in 



341 



10 



9 " 



2 



u. 
o 
o 
O 



6" 



5" 



4-H 



APR 87 




OCT 87 





SEEP 



10 15 

SAMPLING SITE NUMBER 



20 



25 



30 

OUTLET 



Figure 5. --Viable counts of microbes on agar water medium by location in the 
wetland expressed as the log of colony forming units per mL of sediment 
(CFU/mL). Sampling site 1 corresponds to the seep; 29 corresponds to the 
outlet. 



association with bales of hay placed in the 
wetland to alleviate channelization. 



DISCUSSION 

Given the relatively high pH of this mine 
drainage, iron is expected to come out of 
solution to some degree and settle without the 
presence of a wetland. In addition, the 
elevated pH probably enhances the biological 
responses to metals at the cell surface 
(Campbell and Stokes 1985). Thus, the seep pH 
of 6.0, which is relatively unaltered by the 
wetland, positively influences the wetland with 
respect to iron. At this point, it is 
impossible to discriminate between the pH effect 
on iron solubility and the pH effect on 
biological surfaces and cation exchange. 
However, the effect of wetland installation on 
iron removal is significant: iron removal 2 
years after wetland construction is up 80%-90% 
over pre-wetland iron removal. Why is this the 
case? There are a few conspicuous differences 
between pre- and post-wetland sites, and several 
wetland factors to consider regarding the 
wetland 's apparent effectiveness: 

(1) The area of water impoundment has been 
increased by wetland installation. As a result 
the amount of water exposed to the air is much 
greater. 



(2) The retention time of the water was 
probably increased by wetland installation. 
Therefore, any biologically assisted or 
unassisted reactions leading to iron 
precipitation have a longer period in which to 
occur. 

(3) Biological activity in the form of 
microbial, algal, and cattail-cutgrass 
populations is much greater following wetland 
construction and vegetation cover increased 
considerably from the first to the second 
summer. This activity was promoted principally 
by the addition of organically rich compost 
growth medium for the cattails. Cattails are 
highly productive and thereby return a 
considerable amount of organic material to the 
system via winter die-off of stems and leaves. 
Any effect of the substrate apart from 
biological activity cannot account for all of 
the iron removal. During the first winter 
(1985-6) following wetland construction little 
else but spent mushroom compost was in the 
wetland; the winter efficiency was 28%, 
approximating pre-wetland efficiency. However, 
during the second winter (1986-7) efficiency 
increased to 51%. This difference is probably 
attributable to biological activity. 

(4) Cattail roots and rhizomes are directly 
absorbing or precipitating some of the iron. 
Roots from cattails at Simco #4 contained up to 



342 



5 times as much iron as cattail roots from a 
control site. Iron precipitates are also 
loosely associated with cattail roots and 
rhizomes. The cattail rhizosphere may be a 
conducive environment for aerobic microbes such 
as species of Thiobacillus that oxidize iron. 
Further study will determine whether the 
distribution of iron in the substrate of a 
wetland receiving mine drainage is related to 
proximity to roots. 

(5) Wetland Cell 3 was the most effective 
portion of the constructed wetland in improving 
the water quality. It is the largest cell, but 
the degree of water improvement is greater than 
can be accounted by size difference. Other 
characteristics of Cell 3 include the highest 
total flowering plant cover, the greatest 
density of cattail stems, and cutgrass coverage 
exceeding the other two cells. Microbial 
populations were no greater in Cell 3 than in 
the other cells. Cell 3. receives water with the 
lowest level of iron, and thus its higher 
efficiency may, alternatively, reflect (a) a 
water pretreatment effect of previous cells, or 
(b) that iron is more easily removed by a 
wetland at lower incoming levels. 

(6) The diversion ditches and hay bale 
dikes installed in the fall of 1987 had an 
immediate effect on iron removal (increased), pH 
(decreased), manganese (increased), and acidity 
(decreased). It is likely that hay, much like 
sawdust, encourages anaerobic sulphate-reducing 
bacteria that precipitate FeS - FeS 2 (Tuttle 

et al. 1969). On the downstream side of hay 
bales conspicuous areas of black water were 
observed, consistent with this hypothesis. 

(7) The roles of algae in this wetland are 
not yet understood, but their ubiquitous 
presence, both in the floating iron 
surfactant/f locculant (primarily Oscillatoria ) 
and at the substrate-water interface, deserves 
further study. 

(8) Little is known about heterotrophic 
bacteria in acidic mine drainage. However, they 
have been shown to be involved in ferric iron 
reduction and in iron deposition in freshwater 
lakes (Jones 1986). Aerobic and facultatively 
anaerobic bacteria are thought to contribute to 
iron and possibly manganese removal primarily 
through reduction of these metallic species. 
Ferric iron could be reduced to iron oxide 
(FeOOH), which is believed to be the reactant 
with HS~ that leads through mackinawite to 
greigite and finally to pyrite (see Jones 1986 
and references therein). The HS~ is produced 
anaerobically by several species of bacteria. 
Anaerobic bacteria have not yet been counted in 
the wetland, but as noted previously, the 
products of sulfate reduction by bacteria were 
obvious at each of the visits to the wetland. 
The oxidation of iron was evident in the 
wetland, as indicated by the deposition of 
Fe(0H)3 and the formation of yellow-boy. This 
was especially evident around the shallow 
outlets and inlets of the seep and each cell. 
These areas were undoubtedly the most aerobic 
zones in the wetland. Iron-oxidizing bacteria 
of the genus Thiobacillus were found throughout 
the wetland. 



(9) The winter efficiency of wetlands 
receiving mine drainage has frequently been 
questioned. Data from Simco /M indicate only a 
slight reduction in iron-removal efficiency 
between summer and winter (56% to 51%). This 
result is encouraging, but the water quality and 
other site-specific conditions prohibit 
generalizations at this point. 



ACKNOWLEDGEMENTS. This research was made 
possible by support from American Electric Power 
Company (Simco, Inc., contract C-6900), Peabody 
Coal Company, and the Ben Franklin Partnership 
Program. Special thanks to T. Romanowski, B. 
Will, D. Preston, W. Wenerick, E. DeVeau, P. 
Phillips and C. Wilson for their assistance. 



LITERATURE CITED 

Adriano, D. C, R. R. Sharitz, T. G. Ciravalo, 
C. Luvall and S. A. Harding. 1984. Growth 
and mineral nutrition of cattails 
inhabiting a thermally-graded South 
Carolina, USA, reservoir. 2. The 
micronutrients. J. Plant Nutr. 
7:1699-1716. 

Allen, S. E., H. M. Grimshaw and A. P. Rowland. 
1986. Chemical Analysis, pp. 285-344. In: 
Methods in Plant Ecology, 2nd ed. (P. D. 
Moore and S. B. Chapman, eds.), Blackwell 
Scientific Publications, London. 

American Public Health Association. 1980. 

Standard methods for the examination of 
water and wastewater. 15th edition. 
American Public Health Association, 
Washington, D.C. 

Balkwill, D. L. 1977. The natural attachment 
of bacteria to soil. Ph.D. dissertation. 
The Pennsylvania State University, 
University Park, PA. 

Balkwill, D. L. and W. C. Ghiorse. 1985. 

Characterization of subsurface bacteria 
associated with two shallow aquifers in 
Oklahoma. Appl. Environ. Micro. 50:580-588. 

Brooks, R. P., D. E. Samuel and J. B. Hill 

(eds.) 1985. Wetlands and Water Management 
on Mined Lands. The Pennsylvania State 
University, University Park, PA. 

Campbell, P. G. C. and P. M. Stokes. 1985. 
Acidification and toxicity of metals to 
aquatic biota. Can. J. Fisheries and 
Aquatic Sci. 42:2034-2049. 

Daubenmire, R. F. 1970. Steppe vegetation of 
Washington. Wash. Agric. Expt. Sta. Tech. 
Bull. 62. 

Jones, J. G. 1986. Iron transformation by 

freshwater bacteria. Adv. Microbiol. Ecol. 
9:149-185. 



343 



Kleinmann, R. L. P. and M. A. Girts. 1987. Acid 
mine water treatment in wetlands: An 
overview of an emergent technology, pp. 
255-261. In: Aquatic Plants for Water 
Treatment and Resource Recovery. K. R. 
Reddy and W. H. Smith (eds.), Magnolia 
Publishing Inc., Orlando, FL. 

Macfie, S. M. and A. A. Crowder. 1987. Soil 

factors influencing ferric hydroxide plaque 
formation on roots of Typha latifolia L. 
Plant and Soil 102:177-184. 

McHerrcn, L. E. 1986. Removal of Iron and 

Manganese from Mine Drainage by a Wetland: 
Seasonal Effects. M.S. Thesis, The 
Pennsylvania State University, University 
Park, PA. 

Stone, R. W. 1984. The presence of iron- and 

manganese-oxidizing bacteria in natural and 
simulated bogs. pp. 30-36. In: Treatment 
of Mine Drainage by Wetlands. J. E. Burr is 
(ed.), The Pennsylvania State University. 

Tuttle, J. H., P. R. Dugan and C. I. Randies. 

1969. Microbial sulfate reduction and its 
potential utility as an acid mine water 
pollution abatement procedure. Appl. 
Micro. 17:297-302. 

Wichlacz, P. L. 1980. Acidophilic 

Heterotrophic Bacteria from Acid Mine 
Drainages in Central Pennsylvania. M.S. 
Thesis. The Pennsylvania State University, 
University Park, PA. 

Wieder, R. K. , G. E. Lang and A. E. Whitehouse. 
1982. Modification of acid mine drainage 
in a fresh water wetland, pp. 38-62. In : 
Proceedings, Third WV Surface Mine 
Drainage Task Force Symposium, WV Surface 
Mine Drainage Task Force, Charleston, WV. 



344 



TREATMENT OF DISCHARGE FROM A HIGH ELEVATION METAL MINE 
IN THE COLORADO ROCKIES USING AN EXISTING WETLAND 



J.C. Emerick, W.W. Huskie, and D.J. Cooper 1 



Environmental Sciences Department, 
Colorado School of Mines, 
Golden, CO. 



Abstract. During the summer of 1986, discharge from 
the Pennsylvania Mine was diverted into a natural 
sedge wetland in an experiment to assess the metal 
removal capability of a wetland type common to the 
higher elevations of the Rocky Mountains. The 
Pennsylvania Mine is an abandoned metal mine located 
at an elevation of 3355 m in the Peru Creek basin of 
Central Colorado. Surface discharge from the mine 
averages 380 L/min with a mean pH of 3.6 and average 
concentrations (in mg/L; n = 6) of the following 
metals: Fe, 50; Zn, 25; Mn, 20; Al, 15; Cu, 6; Sr, 
0.7; Cd, 0.1; Ni, 0.1; and Pb, 0.1. During the 
course of the investigation it was determined that 
the wetland was heavily contaminated with metals 
prior to the experiment, apparently coming from 
metal-laden surface and colluvial waters. Soil metal 
concentrations (mg/kg; n = 12) averaged: 23,195 for 
Al; 100,899 for Fe; 159 for Mn; 3876 for Cu; 3713 for 
Zn; 3.17 for Ni ; 42.2 for Cd; and 921 for Pb. Total 
soil metals per se was not correlated with the 
qualitative health of plants in the wetland, but high 
concentrations of copper, zinc, nickel and cadmium 
were associated with poor plant health. Because of 
the low hydraulic conductivity of the peat soils 
(10 -3 to 10" 4 cm/sec) the ability of the fen to 
accommodate drainage from the adit was limited and 
the experiment was terminated. However, the study 
demonstrated that the plant species present (mainly 
Carex aquatilis ) have a high tolerance for metals and 
low pH and thus have good potential for use in 
constructed wetland treatment systems in the Rocky 
Mountains. 



x Paper presented at the annual meeting 
of the American Society for Surface 
Mining and Reclamation 



345 



INTRODUCTION 

Mine drainage and leachate from 
mine spoil and mill tailings are 
serious water guality problems in some 
parts of Colorado. Eight Superfund 
sites are focused on the clean-up of 
wastes from decades of mining 
activity, and a recent study (Lewis et 
al. 1987) estimated that out of a 
total of 29,766 km (18,500 mi) of 
streams and rivers in the state, over 
6200 km (3000 mi) have dissolved 
metal concentrations that 
exceed basic standards for aguatic 
life, agricultural use, or domestic 
water supply. Approximately 2,896 km 
(1,400 mi) of that mileage is 
considered to be a direct result of 
pollution from inactive mining 
operations. These problems have 
spawned an interest by Colorado 
agencies to investigate passive 
treatment methods for mine drainage 
abatement. While such methods 
commonly are used in coal mining 
regions of the eastern U.S., colder 
climates, rougher terrain, and greater 
amounts of heavy metals in the Rocky 
Mountain mineral belt have impeded 
direct transfer of this technology to 
many western areas. 

This paper reports the results of 
one such investigation conducted 
during the summer and fall of 1986 
when a natural high-elevation wetland 
was used in an attempt to treat 
discharge from an inactive mine. 

While the treatment attempt met 
with poor success, the project 
provided an opportunity to assess 
metal accumulation in soils and native 
plants, and tolerance of the plants to 
high metal concentrations. Biological 
and geochemical processes that are 
responsible for metal retention in 
mountain wetlands are poorly 
documented. The work described here 
is being conducted with the aim of 
increasing the knowledge of these 
processes, thus leading to the 
development of criteria for the 
feasibility and design of constructed 
wetlands to treat mine drainage at 
high elevations in the Rocky 
Mountains . 



SITE DESCRIPTION 

The Pennsylvania Mine, located at 
an elevation of 3355 m (11,000 ft) 
near Peru Creek in the Front Range of 
central Colorado, is an inactive gold 
and silver mine with thousands of 
meters and many levels of underground 



workings. Surface discharge from the 
mine averages 3 80 L/min with a mean pH 
of 3.6 and mean concentrations (in 
mg/L; n = 6) of the following metals: 
Fe, 50; Zn, 25; Mn, 20; Al, 15; Cu, 6 ; 
Sr, 0.7; Cd, 0.1; Ni, 0.1; and Pb, 
0.1. The mine drainage seriously 
impacts Peru Creek, and in many ways 
is typical of mine drainage problems 
in the Colorado mineral belt. Prior 
to this study, the discharge from the 
Pennsylvania mine flowed in a ditch 
from the adit to Peru Creek, dropping 
approximately 25 m over a distance of 
150 m. 

During a portion of this study, 
mine drainage was diverted into a 1.6 
ha wetland located approximately 180 m 
from the mine and a few meters above 
the Peru Creek channel. The wetland 
is technically a fen, since it is 
underlain by highly organic soils and 
is receiving nutrients and water from 
sources other than precipitation 
(Daubenmire, 1968). It is dominated 
by a near monoculture of water sedge 
( Carex aguatilis ) interspersed with 
patches of bog birch ( Betula 
glandulosa ) growing on many of the 
larger hummocks. The soil surface is 
covered by 1 to 5 cm of water over 
much of the wetland throughout the 
growing season. Soils are 
characterized by accumulations of peat 
to depths of up to 2 m. The upper 10 
cm of soil averages 41 percent organic 
matter, with the texture of the 
inorganic fraction typically loam or 
clay loam. Over much of the wetland 
the upper 3-4 cm of soil are oxidized 
and are characterized by reddish iron 
oxyhydroxides, and below this depth 
the soil consists of gray to black 
reduced sediments with an occasional 
strong smell of hydrogen sulfide. 
Ruins of several mine buildings 
including a small mill are located 
just above the wetland, and mill 
tailings cover the surface of the 
western third of the wetland. The 
majority of plant and soil samples 
taken for characterization of 
"natural" conditions were from the 
eastern half of the wetland where 
impacts from mining activities 
appeared to be minimal . However , 
several shallow wells augured into the 
soil in that portion of the wetland 
indicated lenses of buried sandy 
material that might have been mill 
tailings . 

DESCRIPTION OF THE STUDY 

A mine drainage diversion system 
was constructed by the Colorado 
Division of Mined Land Reclamation 



346 



during the fall of 1985. This system 
consisted of a PVC pipe that carried 
the mine water from the adit to a 150 
m perforated leach line, which was 
buried at a depth of 1.5 m along the 
upper margin of the wetland. In 
addition, ten 5-cm diameter wells were 
installed in the organic soil layer of 
the wetland to monitor metal 
concentrations in the interstitial 
water. The wells were grouped into 
three parallel transects approx- 
imately 30 m apart and roughly 
perpendicular to the expected flow of 
water through the wetland. The upper 
transect closest to the leach line was 
comprised of three wells, the middle 
transect consisted of four wells, and 
the remaining three wells were in the 
lower transect. 

A sampling program was carried out 
during the summer and fall of 1986 and 
1987 to monitor metals in Peru Creek 
and interstitial water in the wetland, 
as well as metal concentrations in 
soils and plants. This paper 
summarizes some of the results of that 
sampling program, chiefly from the 
well transects mentioned above, and 
from six intensive study sites in the 
wetland where metals in soils and 
plants were monitored. The leach line 
system was activated from August 25, 
1986, to October 29, 1986. Three 
series of water samples were collected 
from the wetland wells prior to the 
diversion of mine water, and one 
series was collected 30 days following 
diversion. During 1987, metal 
sampling continued in soils and 
plants, although the leach 
line system was not reactivated. 

All water and soil samples were 
analyzed by the Colorado State 
University Soil Testing Laboratory. 
Water samples were analyzed by 
inductively coupled plasma emission 
spectrometry (ICP). Soil metals were 
determined by nitric-perchloric acid 
digestion followed by ICP analysis, 
and cation exchange capacities were 
measured by displacement using 
ammonium chloride as the saturating 
salt, using methods based on Page et 
al (1982). Metals in plant samples 
were extracted with sulfuric acid and 
then analyzed by ICP. 

RESULTS 

Diversion of Mine Drainage 

Introduction of the .mine drainage 
into the wetland via the leach line 
system was partially successful. 



Average concentrations of several 
metals in the upper transect of ground 
water monitoring wells were 
significantly higher than before the 
system was activated (Table 1), 
indicating percolation of mine 
drainage into the wetland. However, 
the relatively low hydraulic 
conductivity of the peat soils 
(calculated using Darcy's law to be in 
the range of 10" 3 to 10" 4 cm/sec) 
did not permit the entire flow from 
the adit to pass into the wetland, nor 
did the seepage from the leach line 
appear to pass beyond the upper well 
transect. The average metal 
concentrations in the two lower well 
transects did not significantly change 
from the time the system was activated 
until it was turned off and monitoring 
was terminated for the winter. While 
the diversion system was in operation, 
excess mine drainage that could not be 
accommodated by the peat soils leaked 
to the ground surface near the leach 
line and had to be routed around the 
wetland with a small interception 
ditch. The diversion system was not 
reactivated during 1987. 

Metals in Plants and Soils 

A preliminary survey of the 
wetland indicated that while the 
herbaceous cover consisted almost 
entirely of one sedge species ( Carex 
aquatilis ) , .plant growth and health 
was not uniform. More intensive 
vegetation measurements indicated a 
range in cover from 5 to 56 percent, 
and peak season standing crops ranging 
from 18 to 721 g/m 2 (oven-dried 
weight) of aboveground biomass. Areas 
of low cover and biomass were 
characterized by spindly, chlorotic 
plants, whereas sites of high cover 
and biomass typically had large, 
robust, green plants. 

Surface water samples collected 
prior to activation of the leach line 
contained metal concentrations that, 
for some elements, mirrored 
concentrations found in the mine 
discharge. Table 2 shows data for 
five sites on the wetland, and 
illustrates that poor health was not nec- 
essarily associated with higher metal 
concentrations. Sites of poor health 
were usually found in areas that were 
poorly drained or in depressions 
between hummocks . 

Mean concentrations of metals in 
soils and in sedge leaves are shown in 
Table 3. Carex aquatilis apparently 
is able to tolerate high concen- 
trations of aluminum and iron, 



347 



Table 1. Mean metal concentrations (mg/L) in groundwater wells before and after 
diversion of mine water. 



Al 



Fe 



Mn 



Cu 



Zn 



Ni 



Cd 



Pb 



Upper 


Transect 


(3 Wells) 














Before 


(n=9 

S.D. : 


11.85 
1.02 


0.61 
0.79 


6.45 
1.35 


1.55 
0.17 


7.22 
0.73 


0.06 
0.01 


0.04 
0.01 


0.07 
0.02 


After 


(n=3) 
S.D. : 


23.03* 
2.64 


2.33* 

1.37 


9.44 
6.14 


1.58 
1.13 


13.67 
2.15 


* 0.10* 
0.01 


0.08* 
0.01 


0.11* 
0.00 


Middle 


Transed 


z (4 Wells) 














Before 


(n=12 
S.D. : 


3.70 
3.76 


12.56 
11.13 


5.23 
2.37 


0.02 
0.01 


0.94 
1.42 


0.03 
0.02 


0.01 
0.00 


0.06 
0.02 


After 


(n=4) 
S . D . : 


4.53 
5.86 


16.10 
2.45 


4.97 
2.40 


0.05# 
0.04 


0.86 
1.11 


0.33 
0.03 


0.01 
0.00 


0.09 
0.08 


Lower ' 


rransect 


(3 Wells) 














Before 


(n=9) 
S.D. : 


16.28 
17.79 


5.92 
5.31 


9.17 
4.80 


0.10 
0.15 


8.55 
9.44 


0.08 
0.05 


0.05 
0.06 


0.12 
0.13 


After 


(n=3) 

S.D. : 


11.80 
9.64 


9.13 
4.44 


7.14 
3.46 


0.03 
0.03 


4.36 
5.83 


0.06 
0.04 


0.01 
0.00 


0.15 
0.12 


Note: 


(*) significant 
(#) significant 


increase at 99% 
increase at 95% 


confidence 
confidence 


level 
level 







S.D. is standard deviation 



Table 2. Metal concentrations and pH of five surface water samples collected from 
the Pennsylvania mine wetland on July 23, 1986, with corresponding plant 
cover and biomass measured at the sample locations. 



Plants 



Metals (mg/L) 



pH 



% Cover 


g/m : 


; 


Health 


Al 


Fe 


Mn 


Cu 


Zn 


Ni 


Cd 


"Pb 




8.4 


109 


9 


P 


16.4 


13.0 


2.44 


1.74 


0.10 


0.02 


0.01 


0.05 


3-3 


8.8 


42 


6 


P 


14.0 


0.13 


10.3 


1.88 


10. 9 


0.07 


0.07 


0.06 


3.6 


18.0 


68 


3 


I 


15.7 


0.22 


12.4 


1.70 


13.9 


0.09 


0.07 


0.09 


'3.2 


28.0 


492 





I-R 


12.6 


0.56 


8.81 


2.23 


11.2 


0.07 


0.10 


0.05 


3.6 


44.0 


536 





I-R 


12.7 


0.50 


8.11 


2.00 


8.55 


0.06 


0.06 


0.06 


3.8 



Note: P denotes poor, I denotes intermediate, and R denotes robust 



348 



Table 3. Mean concentrations of metals (mg/kg) from samples 
of soils and Carex aquatilis leaves collected 
during the summer of 1987 (n = 2 for each site). 



Plant 
Health 



Al 



Fe 



Mn 



Cu 



Zn 



Ni 



Cd 



Pb 



SOILS 



Poor ( 2 sites) 





14285 


6323 


143 


8298 


10228 


9.5 


119.8 


1037 


S.D. : 


4035 


3900 


17.9 


1953 


3392 


2.9 


66.0 


855 


Intermediate ( 3 


sites) 


















26517 


178167 


195 


2017 


370 


<1 


4.0 


750 


S.D. : 


6222 


58687 


31.9 


1544 


176 


0.0 


1.8 


126 


Robust (1 site) 




















31050 


58250 


85 


612 


717 


<1 


3.0 


1201 


S.D. : 


3950 


37850 


25 
PLANTS 


395 


199 


0.0 


0.0 


171 



Poor 





92.8 


372 


696 


30.1 


351 


5.0 


0.75 


5. 


S.D. : 


23.3 


243 


26 


3.3 


77 


1.2 


0.25 


0. 


diate 


114.8 


516 


621 


80.7 


427 


9.5 


1.27 


<5 


S.D. : 


40.7 


134 


70 


26.8 


78 


1.5 


0.47 


0.0 



Robust 



S.D. 



2.5 


160 


314 


28.5 


257 


3.5 


0.80 


6.0 


3.5 


33 


38 


5.5 


13 


0.5 


0.20 


1.0 



Note: S.D. is standard deviation 



and under the conditions observed so 
far at the wetland, we have seen no 
significant correlation between 
amounts of these metals in the soils 
and corresponding plant health. 
However, high copper concentrations 
in the soil frequently is associated 
with poor growth, and to a lesser 
extent zinc, cadmium, and nickel have 



shown a similar trend. The highest 
values of these metals were always 
associated with sites of poor health, 
while the lowest concentrations were 
found on the healthiest sites. We 
observed no correlation between plant 
growth and total soil metals, even 
when total metal concentrations at 
some sites exceeded 300,000 mg/kg. 



349 



An important growth factor is the 
nutrient status of the plants. 
Metals can inhibit nutrient uptake 
and metabolism, causing nutrient 
deficiency symptoms, and mountain 
soils are often low in certain 
nutrients (Johnson and Cline, 1965). 
We examined soil calcium, nitrogen, 
phosphorus, and potassium, and found 
that of these four nutrients, only 
soil potassium concentrations showed 
a significant correlation with plant 
cover (Figure 1). 

Organic matter in the soils 
ranged from 5.8 percent to 79.1 
percent, with mean percentages of 
59.7, 35.8, and 16.5 for the poor, 
intermediate, and robust sites, 
respectively. Cation exchange 
capacities (CEC) were relatively low, 
with a range of 3 to 24 meg/lOOg, and 
a mean of 12.7. As would be 
expected, CEC values appeared to be 
closely related to the organic matter 
content of the soils. Metals 
extracted (displaced) during the CEC 
determinations of the soil samples 
represented from 1 to 12 percent of 
the total metals in the samples. 

DISCUSSION AND CONCLUSIONS 

Results of the mine drainage 
diversion study were inconclusive. 
The system was not in operation for a 
long enough period of time to permit 
percolation of the mine water 
completely through the wetland. 
However, given the relatively high 
pre-existing metal concentrations in 
the groundwater and soils of the 
wetland, it is questionable whether 
the system would have had much of a 
beneficial impact on water quality. 

Analyses of the plants, soils, 
and surface and subsurface waters 
subsequently have verified that metal 
concentrations throughout the wetland 
are high. The source of the metals 
is apparently from surface and 
colluvial waters flowing from the 
slope above the wetland, probably 
passing through a portion of the 
waste rock dump of the mine. Because 
of the relatively thick peat 
accumulations, the age of the wetland 
is probably several thousand years 
old, and it is possible that it was 
receiving metals before mining 
activities began nearly a hundred 
years ago. We have been conducting 
studies on other wetlands in this 
portion of the Colorado mineral belt 
and have found several sites that are 
receiving high metal loads of natural 
origin. 



High metals concentrations per se 
do not seem to limit the growth of 
Carex aquatilis in the wetland. 
However, poor growth, characterized 
by low cover and biomass, spindly 
shoots, and chlorotic leaves appears 
to be associated with high soil 
concentrations of copper, zinc, 
nickel, and cadmium, as well as 
with low potassium concentrations. 
Perhaps the most significant factor 
adversely affecting plant health is 
poor drainage, judging from visual 
observations at the site. The 
healthiest plants generally are 
living on microtopographically high 
spots or on slopes where drainage is 
obviously better. These sites are 
also associated with lower 
percentages of soil organic matter. 
It is also possible that greater 
amounts of soil organic matter on 
sites of poorer health contribute to 
the formation of metal-organic 
complexes that may increase the 
availability of some metal species to 
the plants (Thornton, 1986). 
Clearly, the uptake of metals by 
Carex aquatilis and the factors 
leading to toxic effects are complex 
issues that will require a more 
detailed study than was possible 
here. 

It is apparent that Carex 
aguatilis is tolerant of high metal 
concentrations and low pH and thus 
would be a likely candidate for use 
as a plant species in wetlands 
constructed at high elevations. 
Success of transplanting and rates of 
establishment remain to be determined 
for this species. 

Metal absorption and retention by 
the soils from cation exchange 
appears to be relatively low and less 
important than other processes. 
Microbial uptake has not been studied 
at this site but presumably would 
play a role in metal retention by the 
wetland (Tuttle et al, 1969). Given 
the cold climate and relatively short 
growing season at mountain sites such 
as the Pennsylvania mine, it is 
likely that many of the 
biogeochemical processes that are 
responsible for metal accumulation in 
wetlands would be slower than at 
lower elevations. While this study 
has provided some insight into metal 
accumulation at one high-elevation 
wetland, a much greater research 
effort is needed before the 
feasibility of using a constructed 
wetland approach in such locations 
can be determined. 



350 




20 40 

PERCENT PLANT COVER 



Figure 1. The relationship between 
soil potassium concen- 
trations and percentage of 
plant cover on the Pennsyl- 
vania Mine wetland. 



ACKNOWLEDGMENTS 

Support for this research is being 
received from the U.S. Environmental 
Protection Agency, Region VIII, and 
Colorado Department of Natural 
Resources, Division of Mined Land 
Reclamation. 



eo 



Johnson, D.D. , and Cline, A.J. 1965. 
Colorado Mountain Soils . Colo. 
Agric. Exper. Sta. Sci. Series 
Paper No. 996:233-281. 

Lewis, W.S., Huskie, W.W. , Woldow, C. 
and Emerick, J.C. 1987. An 
evaluation of mining related 
metals pollution in Colorado 
streams . Draft Report to the 
Colorado Department of Natural 
Resources Division of Mined Land 
Reclamation, September, 1987. 

Page, A.L., Miller, R.H. , and Keeney, 
D.R. [Eds.] 1982. Methods of 
Soil Analysis , 2nd Ed. Amer. Soc. 
Agronomy and Soil Sci Amer., 
Madison, WI . 

Thornton, I. 1986. Soil and plant 
factors that influence element 
availability and uptake: 
implications for geochemical 
prospecting, in Carlisle, D. , 
Berry, W.L., Kaplan, I.R., and 
Watterson, J.R. (Eds.) Mineral 



Exploration: Biological Systems 
and Organic Matter , Rubey Volume 
V. Prentice-Hall, N.J. 

Tuttle, J.H. , Dugan, P.R., and 

Randies, C.I. 1969. Microbial 
sulfate reduction and its 
potential utility as an acid mine 
water pollution abatement 
procedure. Applied microbiology 
17:297-302. 



LITERATURE CITED 



Daubenmire, R.F. 
Communities : 



1968. Plant 

A Textbook of Plant 



Synecology . Harper and Row, N.Y. 



351 



THE TRACY WETLANDS: A CASE STUDY OF 
TWO PASSIVE MINE DRAINAGE TREATMENT SYSTEMS IN MONTANA 



Michael T. Hiel and Francis J. Kerins, Jr. 



Abstract. — Two man-made wetland systems were con- 
structed in Montana to experiment with and assess appli- 
cability of Passive Mine Drainage Treatment technologies to 
the state. The wetland systems, designated the Large Tracy 
Wetland and the Small Tracy Wetland treat flows of .95 and 
.15 liters per second, respectively. Both wetlands were 
constructed during the summer of 1986 and utilize a peat 
substrate planted predominantly with Typha latifolia for 
metals removal and limestone gravel and aeration structures 
for pH buffering. The large wetland has approximately 418 
square meters of surface area. The inflow to this system has 
metals concentrations in excess of 280 mg/L Al and 1.5 mg/L 
Mn. The small wetland has approximately 111 square meters 
of surface area. The inflow to this system has metals 
concentrations in excess of 143 mg/L Fe, 46 mg/L Al and 
mg/L Mn. Construction costs for the Large and Small Wetland 
were $67/m 2 and $140/m 2 , respectively. Both wetlands were 
relatively ineffective in improving the water quality of the 
acid mine drainage. Low system retention times and minimal 
contact between the peat and acid mine drainage are primary 
reasons for the ineffectiveness of the systems. 



INTRODUCTION 

The Abandoned Mine Reclamation (AMR) 
Bureau of the Montana Department of State 
Lands (DSL) has identified more than 50 sites 
in Montana where acid mine drainage from 
abandoned mining facilities has significantly 
impacted the local environment. Passive mine 
drainage treatment techniques provide a 
potentially useful and economical means for 
mitigation of acid mine drainage impacts at many 
of these sites. 



Michael T. Hiel is Reclamation Specialist, 
Mont. Dept. of State Lands, Helena, MT, and 
Francis J. Kerins, Jr. is a Mining Engineer, 
Robert Peccia & Associates, Helena, MT. 



Two man-made wetland treatment systems 
were designed and constructed in Montana to 
treat acid mine drainage from abandoned coal 
mines. These wetlands, designated the Large 
Tracy Wetlands and the Small Tracy Wetlands, 
are located near the town of Tracy in the Sand 
Coulee drainage ten miles southeast of Great 
Falls in Cascade County. The wetlands were 
constructed during July and August of 1986. 

This report presents design criteria and 
construction costs for both wetland systems. In 
addition, inflow and outflow water quality data, 
soil chemistry data, and plant tissue chemistry 
data are presented. 

SITE DESCRIPTION 

The Large Wetland treats a flow of .50 to 
.95 liters per second (11,520 to 22,000 gpd). 
This flow has a pH of 2.7, and total iron, total 



352 



aluminum, total manganese, and sulfate concen- 
trations of 284 mg/L, 178 mg/L, 1.51 mg/L and 
2618 mg/L, respectively. 

The Large Wetland system includes a peat 
layer planted with cattails and sedges for metals 
removal, a limestone gravel channel for the 
neutralization of mineral acidity, and several 
aeration structures. A limestone-soil mix was 
used as a substrate throughout the system and 
in the construction of baffles to provide 
sinuosity . 

The Small Wetland treats a flow of .38 to 
.50 liters per second (8,500 to 12,000 gpd). 
This flow has a pH of 3.1, and total iron, total 
aluminum, total manganese and sulfate concen- 
trations of 148 mg/L, 46.7 mg/L, 1.2 mg/L, and 
1560 mg/L, respectively. 

The Small Wetland includes a peat layer, a 
limestone gravel channel, aeration structures, 
and a limestone-soil substrate. This wetland was 
also planted with cattails and sedges. 

SITE LOCATION, HISTORY AND HYDROLOGY 

The Great Falls area is a semiarid envi- 
ronment with a total precipitation of 38 cm (15 
inches) per year. Seventy percent (70%) of the 
annual precipitation occurs during the months 
of April through September. The annual mean 
temperature is 7.2°C (45°F), with extremes 
from 37.8°C to -31.7°C (100°F to -25°F). Temp- 
erature and rainfall have a significant effect 
upon hydrology and soil characteristics. 

The Sand Coulee drainage, consisting of 
approximately 500 sq. km (195 sq. mi.) in the 
upper Missouri River basin, is adversely 
impacted by acid mine drainage (AMD) from 22 
discharging abandoned coal mines. This is the 
largest concentration of abandoned discharging 
coal mines west of the Mississippi River. 
Approximately 230 sq. km (90 sq. mi.) of the 
drainage contains abandoned coal mines dis- 
charging acid mine water. 

The coal in this area occurs within the 
upper part of the Mission Formation (Jurassic) 
and is exposed along outcrops in the valley of 
Sand Coulee Creek and its tributaries. The 
topography is gentle with the exception of 
drainages incised into the plateau leaving a 
dissected and irregular terrain above the 
drainage areas. The Sand Coulee drainage lies 
in the Great Falls-Lewistown Coal Field, which 
is a large deposit of sub-bituminous coal. 
Unlike Eastern Montana Tertiary coal deposits, 
the coal in this area is higher in grade (11,118 
btu/lb), and higher in sulfur content (0.5 - 
5.5%) (Silverman and Harris, 1967). Thickness 
of the coal seam varies from .3 m to 3.6 m (one 
to twelve feet) . 



This coal was the target of the mining 
activity beginning in the 1880's. The last 
large-scale mine closed in 1952. During the 
period from 1885 to 1955 the coal mined from 
the Great Falls-Lewistown Coal Field exceeded 
36 million tons. This amounted to about 23 
percent of the total coal produced in Montana 
during that period. Coal mined from the Great 
Falls-Lewistown area from 1955 to 1965 was less 
than 1 percent of the total coal produced in 
Montana and since 1965 there has been no com- 
mercial coal production (Westech/Hydrometrics, 
1982). 

The geology of the area consists primarily 
of Cretaceous, Jurassic and Mississippian 
sedimentary rocks; the coal occurs in the 
Jurassic rocks. The Cretaceous Kootenai 
Sandstone and Flood Sandstone are aquifers 
most consistently used for water for domestic 
wells and springs in the area. Water from these 
aquifers leaks into the abandoned mines in the 
underlying formations. 

The presence of large, abandoned under- 
ground coal mines in the area has apparently 
produced a large change in the regional ground 
water flow system. The underdraining of the 
basal Kootenai Sandstone aquifer by the 
abandoned mines has diverted the ground water 
flow, which most likely discharged to the Sand 
Coulee drainage prior to mining (Osborne, 
1987). 

The two mine discharges dealt with in this 
project are from the old Pierce Mine. Although 
a relatively small mine, the mine discharges 
have had a significant impact on the adjacent 
agricultural lands. For 40 years the discharges 
had been flowing into prime bottom land 
rendering .08 to .12 sq. km (20 to 30 acres) 
unsuitable for agricultural purposes. The 
problem persisted until even after several 
attempts by the landowner to seal the mine 
opening and reroute the flow of the discharge. 
In the fall of 1984, the Montana Abandoned Mine 
Reclamation Bureau (AMRB) completed a project 
which installed a drainage pipeline system. This 
pipeline collected drainage from both mines into 
a main pipeline and discharged directly into 
Sand Coulee Creek. Subsequent to the wheat 
field drying out, the soils have been amended 
and a wheat crop is once again growing. In 
solving the problem of flooding of the fields 
with acid mine drainage no attempt was made to 
treat the problem of acid mine drainage until 
this project in 1986. 

WETLAND DESIGN 

The sites chosen for construction of the 
experimental wetlands offered several 

advantages, including: 



353 



1. Two different controlled flows of less 
than one liter per second in proximity to each 
other; 

2. Two flows having different pH's and 
metal concentrations; 

3. Easy access to the sites during all 
seasons; and 

4. A topography at the sites which 
provided some freedom in the design of the 
wetlands. 

The larger flow where the Large Wetland 
was constructed has been monitored since 1970. 
The smaller flow where the Small Wetland was 
constructed has been monitored since 1986. 



with some mineral soil content. The minimum 
depth of the peat was .3m (1 ft.). 

5. Avoiding "short-circuiting" of the 
wetland by the formation of flow channels. 

6. Providing for placing cattail sod mats 
over approximately 40 percent of the surface 
area of the wetlands. 

7. Providing a crushed limestone-filled 
channel downstream of the wetland for 
moderation of pH. 

8. Providing aeration structures along the 
limestone channel. 

DESCRIPTION AND CONSTRUCTION 



Design criteria for the wetlands were 
established following communication with people 
involved in the construction of similar systems 
in other states and a review of literature con- 
cerning wetlands construction techniques. The 
principal sources of information were the 
Colorado Inactive Mine Land Program (Holm, 
1986) and a set of notes prepared by Kleinmann 
et al. (1986) for a course entitled, 
"Constructing Wetlands for the Treatment of 
Mine Water." 

The intent of the wetland design was to 
provide a three-stage treatment facility: 
man-made peat wetlands planted with cattails 
and sedges, limestone-filled drainage channels, 
and aeration structures. The peat wetland 
planted with cattails and sedges provides an ion 
exchange facility for the removal of heavy 
metals. The limestone-filled channel provides a 
source of alkalinity for neutralization of the low 
pH drainage. The aeration structures provide a 
means for the exsolution of carbon dioxide to 
the atmosphere thereby reducing the con- 
centration of carbonic acid (Guertin et al. , 
1985). 

The primary criteria used for the design 
of the wetlands included: 

1. Sizing the wetland to allow for 
treatment of the quantity of water expected 
during all seasons and following precipitation 
events. A surface area of 294 m 2 /L/s (200 
ft 2 /gpm) was considered a minimum size 
(Kleinmann, 1986). 



2 . Minimizing water velocities 

maximizing retention time in the system. 



and 



3. Providing water depths varying from 5 
cm to 46 cm (2 in. to 18 in.) in the system. 

4. Providing optimum wetlands soils for 
emergent hydrophytes (such as Typha ) 
composed of decomposed organic matter (peat) 



The Large Wetland was designed as a rec- 
tangular impoundment with the approximate di- 
mensions of 30 meters (100 ft.) by 14 meters 
(45 ft.) (Figure 1). This impoundment was baf- 
fled by a berm extending 28 meters (93 ft.) 



Figure I— Large Wetland 

Outlet Weir 

Aeration Structure 




Inlet Weir 



-Limestone Channel 
PLAN VIEW 




"-Cattail Sod Mat 
— .3 m Peat 
-.45m Limestone Mix 
- Bentonite Liner 
SECTION A- A' 



354 



along the length of the wetland from each side 
of the wetland. This baffling provided sinuosity 
for the flow through the wetlands and pre- 
vented short-circuiting of the flow across the 
wetlands. 

The water level is controlled by a rec- 
tangular notch weir at the outlet. The flow from 
this outlet structure passed over 37 linear 
meters (124 l.f.) of limestone channel and three 
aeration structures before exiting the system. 
Each of the aeration structures had about .45 
meters (1.5 ft.) of fall. 

The Small Wetland was designed as two 
parallel linear impoundments connected at one 
end by an aeration structure, forming essen- 
tially a U-shaped wetland (Figure 2). Each of 
the impoundments is about 18 meters (60 ft.) 
long and 3 meters (10 ft.) wide. The flow from 
the outlet structure passed over approximately 
12 linear meters (40 l.f.) of limestone channel 
and two aeration structures before exiting the 
system . 

The soil underlying both wetlands was 
amended with bentonite to prevent infiltration of 
the water in the impoundments into the ground. 
The base of both wetlands and the berms used 

Figure 2- Small Wetland 

Aeration Structure 




Outlet Weir 



Aeration Structure- 1 
PLAN VIEW Inlet Weir- 



Cattail Sod Mat 




.1 5m Limestone Mix 
Bentonite Liner 



SECTION A- A' 



for baffling in the Large Wetland were made of 
a limestone-soil mix with a minimum depth of .45 
meters (1.5 ft.). A minimum of .3 meters (1 
ft.) of peat was placed over the base material. 
The peat was from an active peat extraction 
operation located approximately 400 kms (250 
miles) southwest of the project site. The 
species of moss, Polytrichum piliferum and 
Polytrichum strictum are common to Montana 
wetland communities and are more often found 
in drier locations. Cattail sod mats measuring 
approximately 3.3 sq. meters (36 sq. ft.) were 
placed on top of the peat, spaced evenly over 
approximately 40 percent of the area of the 
wetland. Sedge mats measuring approximately 
.09 sq. meters (.3 sq. ft.) were placed 
randomly throughout the wetlands. 

The cattail and sedge vegetation sod mats 
were excavated from a wetland source area 
about 250 meters (835 ft.) from the constructed 
wetlands. The mats were excavated using a 
rubber-tired front end loader. The average 
thickness of these mats was about 15 cm. (.5 
ft.). Whole 2.7 meter by 1.2 meter (9 ft. by 4 
ft.) mats were placed in the wetlands. 

In order to minimize the stress on the 
transplanted vegetation uncontaminated water 
was used to partially fill the wetlands prior to 
inundation with mine drainage. The transplanted 
vegetation was allowed to adjust to the new 
environment for one week prior to coming in 
contact with mine drainage. It is believed that 
his procedure helps account for the significant 
"new shoot" growth shown after transplanting 
the Typha . No plant stress has been observed, 
and plant growth has been vigorous. 

MONITORING 

Influent and effluent water quality; metals 
concentrations in Typha roots, rhizomes and 
leaves; and metals concentrations in the peat 
for both systems have been monitored since the 
end of July 1986. Initially, water sampling was 
performed on a biweekly basis, but after the 
first two months a monthly sampling program 
was adopted. Vegetation and peat sampling was 
performed on a monthly basis. 

Two water samples were collected at the 
inlet and outlet of each wetland. One of the 
samples taken from each location was filtered 
and preserved with acid (HNO ). The pair of 
samples from each location was analyzed for 
total and dissolved aluminum, iron and 
manganese. The sulfate concentrations, pH, and 
specific conductance for each pair of samples 
was also determined. The samples were analyzed 
utilizing EPA drinking water methods and ICP 
(Inductively Coupled Argon Plasma) techniques. 

The Typha rhizome, root and leaf samples 
were collected from randomly selected plants in 
the wetlands. After collecting the samples the 



355 



rhizomes and roots were dried and any residual 
peat was carefully removed. Each of the vege- 
tative samples was analyzed for total aluminum, 
iron, manganese and sulfur. The samples were 
processed using nitric acid and hydrogen pe- 
roxide digestion (EPA method 3050). Analysis of 
the samples utilized ICP techniques. 

The peat samples were collected at 
locations near the inlet and outlet of each 
wetland. The samples were collected from the 
upper 15 cm (.5 ft.) of the peat layer. The 
samples were analyzed for a number of con- 
stituents, including total sulfur, pH and total 
and extracted iron, aluminum, and manganese. 
The samples were analyzed utilizing Montana 
Department of State Lands strip mining soil 
analysis methods. 



PERFORMANCE 

Neither the Large Wetland nor the Small 
Wetland has improved the quality of the water 
flow through the system to any degree. Table 1 
shows a summary of the water quality data for 
the inlet and outlet of each of the wetlands. 

The concentrations of iron, aluminum, and 
manganese in the rhizomes, roots and leaves of 
the cattail plants in the two wetlands and the 
cattail source area used as a control are sum- 
marized in Table 2. This data indicates that the 
Typha is concentrating metals in the system. 
However, because the potential uptake of metals 
by Typha is not significant when compared with 
the actual loading of the system, the vegetation 



TABLE 1 
Water Quality Summary 



Large Wetland 



Inlet 



Mean 



Total Al 


178 


Total Fe 


284 


Total Mn 


1.51 


Sulfate 


2618 


pH 


2.7 


Specific 


3349 


Conductivity 





Stan Dev 

17.5 
134.9 
.14 
285 
.27 
307 



Outlet 



Mean 

180 
271 
1.67 
2683 
2.58 
3440 



Stan Dev 

17.9 
110.6 
.23 
201 
.19 
194 







Small 


Wetland 






Inlet 






Outlet 


Mean 


Stan 


Dev 


Mean 


Stan Dev 


46.7 




2.1 


45.7 


2.1 


148.5 




37.2 


94.1 


52.7 


1.2 




0.1 


1.3 


0.3 


1560 




223 


1551 


218 


3.1 




0.3 


2.8 


0.4 


2414 




246 


2559 


278 



NOTES: 1) All units mg/L except for pH (standard units) and Specific Conductance (u mho/cm). 
2) All samples taken between July 10, 1986 and Sept. 30, 1987. Number of samples taken is 21. 



TABLE 2 
Typha latifolia Analysis Summary 







Total Al 






Total Fe 






Total Mn 






Mean 


Stan 


Dev 


Mean 


Stan Dev 


Mean 


Stan 


Dev 


Rhizomes 




















Control 


558 




506 


590 




333 


42 




31 


Large Wetland 


1398 




1268 


4237 




1735 


164 




74 


Small Wetland 


1400 




1478 


4806 




2325 


134 




72 


Roots 




















Control 


5171 




2368 


8093 




7584 


475 




813 


Large Wetland 


7444 




5006 


40884 




25437 


270 




151 


Small Wetland 


4703 




2659 


34609 




29491 


241 




292 


Leaves 




















Control 


139 




100 


198 




95 


275 




263 


Large Wetland 


232 




91 


287 




62 


1760 




1455 


Small Wetland 


243 




101 


286 




134 


1053 




766 



NOTES: 1) All units ug/g. 

2) All samples taken between Dec. 23, 1986 and Sept. 



21, 1987. Number of samples taken is 9. 



356 













TABLE 


3 


















Soil 


Analysis 


Summary 












Control 


Peat 




Large 


Wetland Peat 


Small 


Wetland Peat 




Mean 




Stan Dev 


Mean 






Stan Dev 


Mean 


Stan Dev 


Total Al 


9595 




1665 




12872 






1726.1 


15413 


4606 


Total Fe 


6918 




2358 




29009 






18364 


60839 


68291 


Total Mn 


113 




50 




160.6 






242.7 


98.4 


74.7 



NOTES: 1) All units in ug/g. 

2) Number of samples for control peat is 2. Number of samples for Large and 
Small Wetland peat is 10. 



TABLE 4 
Construction Costs 



Wetlands Construction Costs - 1986 



Large Wetland 
(Dollars) 



Small Wetland 
(Dollars) 



Site Development /Inflow Control 

Facilities 
Wetland Construction 
Materials (in place) 

- Water 

- Bentonite 

- Limestone Gravel 

- Peat 

- Vegetation Sod Mats 
Wetland Flow Control/ Aeration 

Structures 
Total 



$5,510 
5,655 

780 

975 

4,260 

6,760 

440 

3,430 
$27,810 



$4,125 
2,335 

375 

750 

1,470 

3,380 

220 

2,940 
$15,595 



Wetland Surface Area (m 2 ) 
Wetland Surface Area (ft. 2 ) 
Wetland Cost/Area (dollars/m 2 ) 
Wetland Cost /Area (dollars /ft 2 ) 



418 

4,500 

66.53 

6.18 



111 

1,200 

139.89 

13.00 



would provide only a minor contribution to the 
total metal retention of a functioning system. 

The peat has also concentrated metals in 
the system. Total aluminum, iron and manganese 
concentrations for both wetlands and the control 
peat are shown in Table 3. 

Because the removal of the metals in both 
systems was unsuccessful, the limestone gravel 
in the channels downstream of each wetland be- 
came armored in less than two weeks. Although 
no specific tests were made, visual indications 
suggest that metals precipitation and pH buffer- 
ing decreased after the armoring process was 
complete. 

It was also observed that detention times 
in the systems were much shorter than was 
desired. Dye tests showed that the detention 
times were approximately seven hours and three 
hours for the Large and Small Wetlands, res- 
pectively. These low retention times are due 



primarily to the relative impermeableness of the 
peat once it is saturated. 

WETLANDS CONSTRUCTION COSTS 

Construction of the Large and Small 
Wetlands occurred during the months of July 
and August in 1986. Because of the relatively 
good access to the project sites many different 
pieces of equipment were used during construc- 
tion, including bulldozers, backhoes, smooth- 
drum rollers, water trucks, motor graders, 
tractor-trailers, and rubber-tired front end 
loaders. Use of specialized equipment allowed 
for the volumes of material incorporated into the 
system to be easily handled and placed, and for 
work to be completed in less than four weeks. 

Table 4 shows a breakdown of the costs 
for the construction of the Large and Small 
Wetlands. The breakdowns show that material 
costs made up the largest portion of the 
construction costs. The peat was the largest 



357 



cost material item because it was necessary to 
haul it approximately 400 kms (250 miles) to the 
project site. 

CONCLUSION 

Two passive treatment systems constructed 
in Montana during 1986 have been ineffective in 
removing heavy metals from acid mine drainage. 
There seems to be two primary reasons for the 
ineffectiveness of these systems. Firstly, both 
systems were undersized, which resulted in 
very low retention times and limited the contact 
between the AMD and the primary heavy metal 
removal medium, the peat moss. Secondly, the 
systems designs did not force the AMD to flow 
through the peat at any location. The designs 
maximized the length of the flow paths through 
each system, but most of the flow through the 
systems was above the AMD-peat interface. 

Improvements in the performance of these 
systems would probably be realized if the 
retention time was increased and provisions 
were made to force the AMD to flow through the 
peat. Expansions were made to both systems 
during the summer of 1987. These expansions 
significantly increased the size of each system 
and provided for flow of the AMD through the 
peat. A report on the performance of these 
expansions will be made at a future date. 



REFERENCES CITED 

Guertin, D., S. C. Emerick, E. A. Howard, 
1985, Passive Mine Drainage Treatment 
Systems: A Theoretical Assessment and 
Experimental Evaluation. Colorado Mine 
Land Reclamation Division Cooperative 
Agreement 202-317. Colorado School of 
Mines, Golden, CO. 

Holm, David, 1987, Personal communication with 
Michael Hiel during the Spring, 1986. 

Kleinmann, R. L., B. R. Brooks, B. 
Huntsman, and B. Pesavento, 1986, Cons- 
tructing Wetlands for the Treatment of 
Mine Water Course Notes. National Sym- 
posium on Surface Mining Hydrology, Sed- 
imentology and Reclamation, Lexington, 
Ky. 

Osborne, T. J., 1987, Acid Mine Drainage 
Control in the Sand Coulee Creek and Belt 
Creek Watersheds, Montana: Montana 
Bureau of Mines and Geology Interim 
Report (February, 1987). 

Silverman, A. J., and W. L. Harris, 1967, 
Stratigraphy and Economic Geology of the 
Great Falls-Lewistown Coal Field, Central 
Montana: Montana Bureau of Mines and 
Geology Bulletin 56. 

Westech/Hydrometrics, 1982, Abandoned Mine 
Lands Belt-Sand Coulee, Montana: 
Department of State Lands, Montana v. 
Ila, lib, He. 



358 



EFFECTS OF CATTAILS ( TYPHA ) ON 
METAL REMOVAL FROM MINE DRAINAGE 1 



J. C. Sencindiver and D. K. Bhumbla^ 



Abstract. — Natural and constructed wetlands 
have been shown to ameliorate acid mine drainage, 
but mechanisms for metal removal in wetland treat- 
ment systems are not well understood. Therefore, 
seven natural cattail wetlands in Monongalia County 
and five in Preston County, WV, which received 
drainage from surface coal mines were selected for 
study. Aboveground plant biomass in each wetland 
was estimated and plant samples were collected in 
June and September 1 986 to evaluate bioconcentra- 
tion of Fe and Mn . Sediment samples were collected 
to determine pH, Eh, texture, and extractable Fe 
and Mn . Plant biomass of these wetlands was less 
than 50% of biomass reported in the literature for 
wetlands managed for biomass. The average above- 
ground (leaves and stems) uptake of Mn by cattails 
in the spring was 5.37 kg/ha and 5.20 kg/ha for 
Monongalia and Preston Counties, respectively. In 
the fall, Monongalia County cattails accumulated 
11.47 kg/ha and cattails in Preston County accumu- 
lated 6.19 kg/ha. Concentrations of Mn in the 
rhizomes were much lower than concentrations of 
plant tops. Iron uptake by aboveground plant parts 
was very low, less than 4 kg/ha, for both counties 
at both sampling times. Concentrations of Fe in 
the rhizomes were 30 to 100-fold greater than Fe 
concentrations in the plant tops. Calculations 
show that cattails in these wetlands removed less 
than \% of the total Fe added to the wetlands by 
mine drainage. The pe + pH values of the sediment 
where cattails were growing were less than the 
values for areas with no cattails, indicating that 
cattails may lower the redox potential of the wet- 
land enough to precipitate Fe and Mn . 

1 Paper presented at the 1 988 Mine Drainage supported with funds appropriated under 

and Surface Mine Reclamation Conference the Hatch Act and a grant from the West 

sponsored by the American Society for Virginia University Water Research Insti- 

Surface Mining and Reclamation and the tute. 
U.S. Department of the Interior (Bureau of 

Mines and Office of Surface Mining Recla- ' J . C . Sencindiver is an Associate Profes- 

mation and Enforcement), April 17- 22, sor of Soil Science and D.K. Bhumbla is a 

1988, Pittsburgh, PA. Published with the Research Assistant, Division of Plant and 

approval of the Director of the West Vir- Soil Sciences, West Virginia University, 

ginia Agricultural and Forestry Experi- Morgantown, WV. 
ment Station as Scientific Article 
#2102. This research was partially 



359 



INTRODUCTION 



tive 

incr 

the 

and 

1984 

capa 

drai 

tion 

orga 

dema 

rate 

ment 

Patr 

ment 

resp 

cept 



In the 
ways t 
easing 
use of 
Kleinma 
) , beca 
ble of 
nage . 
s and t 
nic mat 
nd of m 
of oxy 
s (Ponn 
ick 197 
s becom 
iration 
ors sue 



search 
o treat 
attenti 
cattail 
nn 1986 
use the 
remov in 
Due to 
he pres 
ter, th 
ost wet 
gen dif 
amperum 
8) . Wi 
e highl 

procee 
h as Fe 



for safe 

acid min 

on has be 

( Tvpha ) 
, Snyder 
se wetlan 
g Fe and 
saturated 
ence of d 
e biochem 
lands far 
fusion in 
a 1972, G 
thout oxy 
y reduced 
ds, using 
, Mn, and 



, cost-e 
e draina 
en focus 
wetlands 
and Ahar 
d system 
Mn from 

soil co 
ecomposa 
ical oxy 

exceeds 
to the s 
ambrell 
gen, the 

as micr 

electro 

so,, 2 -. 



ffec- 
ge 
ed on 

(Girts 
rah 
s are 
the 
ndi- 
ble 
gen 

the 
edi- 
and 

sedi- 
obial 
n ac- 



Cattail plants have morphological 
features such as arenchyema tissues which 
aid in the transport of oxygen from above- 
ground parts to the roots, thereby en- 
abling the plant to grow in highly reduced 
environments (Armstrong 1978, Hutchinson 
1975, Sebacher et al. 1 985) - The direct 
ameliorative effect of cattails is due to 
metal removal by plant uptake from acid 
mine drainage. Metals taken up by the 
plants are subsequently retained in the 
detritus after senescence of the plants. 
The indirect effect of the cattails on 
water quality is related to the fact that 
they modify the chemical and biological 
environments of the sediments. 

Little quantitative data are avail- 
able to provide information about the 
relative importance of direct or indirect 
mechanisms on the water treatment capabil- 
ities of cattail wetlands. Before the 
relative merits of wetland treatment 
schemes can be fully evaluated, it is 
necessary to understand the movement, 
retention, and transformation of metals in 
wetland ecosystems. This information is 
vital for development of design criteria 
for building manmade wetlands for treat- 
ment of acid mine drainage. This investi- 
gation was undertaken a) to determine the 
biomass and elemental composition of cat- 
tails growing in naturally occurring wet- 
lands receiving mine drainage, b) to de- 
termine the seasonal changes in plant 
biomass, plant elemental composition, and 
plant uptake of Fe and Mn , c) to determine 
the temporal changes in extractable metals 
in sediments of cattail wetlands receiving 
mine drainage, and d) to determine the 
effect of the cattail plants on redox 
environment of sediments within the wet- 
lands. 



MATERIALS AND METHODS 

Natural cattail wetlands (major spe- 
cies Typha latif olia ) which received 
drainage from reclaimed surface coal mines 
were selected for study in two counties in 
northern West Virginia. Seven wetlands 
were located in Monongalia County and five 
were located in Preston County. The wet- 



lands ranged in size from 40 m 2 to 1500 
m. Waynesburg coal seam had been mined 
at all of the Monongalia County sites, and 
Upper and/or Lower Freeport coal had been 
mined at the Preston County sites. 

Plant populations in each wetland 
were determined by counting the number of 
plants per square meter. A 1 m x 1 m 
quadrat was randomly placed at five points 
in each wetland. All cattails within the 
quadrat were counted. The total above- 
ground portions (leaves and stems) and 
rhizomes of cattails were sampled near the 
water inflow, near the middle, and near 
the water outflow of each wetland. These 
samples were collected in June 1986 
(Spring) and September 1986 (Fall). The 
plant samples were cleaned with distilled 
water and dried in a forced air oven at 
65°C. Dried samples were ground in a 
Wiley mill having stainless steel blades. 
These samples were sent to Pennsylvania 
State University for Fe and Mn analyses, 
and analyzed on an Inductivity Coupled 
Plasmaspectrograph as described by 
Dahlquist and Knoll (1978). Some plant 
samples were also analyzed at West Vir- 
ginia University for comparison. In this 
procedure (Ganje and Page, 1974), a 1 .0-g 
sample of plant tissue was digested with 
9.0 ml of 3:1 mixture of nitric and per- 
chloric acids in a tall 200-mL beaker. 
The digested material was dissolved and 
brought to volume in a 50-mL volumetric 
flask with 0.1 N HC1. The concentration 
of metals in the solution were determined 
by atomic absorption spectrophotometry. 
Plant uptake of metals was calculated by 
multiplying tissue concentrations by bio- 
mass. Similar data were produced by both 
laboratories, so only one set of data is 
presented. 

Sediment samples were collected from 
the same three points of cattail sampling. 
The sediment samples were collected with 
an auger to a depth of 10 cm from the 
soil-water interface and transferred to 
plastic bags. Soil pH and Eh for redox 
measurements were made in the fresh, wet 
sediments. Eh was measured by using a 
platinum electrode with Ag/AgCl reference 
electrode. Eh readings were noted when the 
drift in Eh was less than 5 mV/min. Sedi- 
ment pH was measured with a standard glass 
electrode. The pe + pH values were calcu- 
lated by dividing Eh by 59.2 and adding pH 
(Lindsay 1979)- After pH and Eh were 
measured the sediments were air-dried and 
passed through a 2-mm sieve. Soil texture 
was determined by the pipette method (Soil 
Survey Staff 1972). Extractable Fe and Mn 
were determined by extracting the sedi- 
ments with 0.1 N HC1 using 1:10 ratio of 
sediment to extracting solution. The 
extracted elements were determined by 
atomic absorption spectrophotometry. 

Additional samples for the study of 
effect of vegetation on pe + pH of the 
sediments were collected from one point 
outside and on the water entry side of 
each wetland. The pe + pH data were cal- 



360 



culated as described above, 



Monthly 
collected fr 
January 1987 
Rauch, Depar 
phy, West Vi 
eral, the Mo 
neutral pH. 
time was 6.3 
Preston Coun 
be somewhat 
consistently 
Water pH at 
below 6.0 an 
4.0. Total 
Preston Coun 
tained sligh 
the Monongal 



water qual 
om November 

(unpublish 
tment of Ge 
rginia Univ 
nongalia mi 

The lowest 

and the hi 
ty, the min 
more acid. 

had pH val 
all other s 
d commonly 
Fe and Mn v 
ty mine wat 
tly more of 
ia County m 



ity sampl 

1985 thr 
ed data, 
ology and 
ersity) . 
ne water 

pH at an 
ghest was 
e water t 

Only one 
ues above 
ites was 
between p 
aried, bu 
er genera 

these me 
ine water 



es were 

ough 

Dr. Henry 

Geogra- 

In gen- 
was near 
y sampling 

8.2. In 
ended to 

site 

6.0. 
generally 
H 3.0 and 
t the 
lly con- 
tals than 



Statistical analysis (ANOVA) was done 
using SAS programs, and significance bet- 
ween counties and seasons was tested at 
alpha levels of 0.05. 



RESULTS AND DISCUSSION 



Primary Biomass Production of Cattails, 



The dir 
effect of ca 
drainage pas 
to primary b 
wetland ecos 
lands in thi 
1 . Dry matt 
for biomass 
puted from p 
yield per pi 
density in M 
signif icantl 
County wetla 



ect o 
ttail 
sing 
iomas 
ystem 
s stu 
er yi 
produ 
lant 
ant ( 
onong 
y gre 
nds . 



r indirect 

wetlands 
through th 
s producti 
Biomass 
dy are pre 
eld was us 
ction, and 
density an 
plant vigo 
alia Count 
ater than 
Cattail g 



amel 
on ac 
em is 
on wi 

data 
sente 
ed as 

it w 
d dry 
r). 
y wet 
in Pr 
rowth 



iorative 

id mine 
related 

thin the 
for wet- 

d in table 
an index 

as corn- 
matter 

Plant 

lands was 

eston 
in wet- 



lands 

cantly 

growth 

sampli 

the me 

in Mon 

times 

Presto 

plant 

times 



of Mo 
more 
in P 
ng ti 
an pe 
ongal 
the d 
n Cou 
vigor 
great 



nonga 
vigo 

resto 

mes. 

r pla 

ia Co 

ry ma 

nty . 
in M 

er th 



lia County 
rous than 
n County w 

During th 
nt dry mat 
unty wetla 
tter yield 

At the fa 
onongalia 
an that of 



was signifi- 
the cattail 
etlands at both 
e spring season 
ter production 
nds was 2.5 

per plant in 
11 sampling, the 
County was 2.2 

Preston County. 



Biomass production was three 
times higher in Monongalia County than in 
Preston County at the spring sampling and 
almost three times higher at the fall 
sampling. Biomass production depended on 
plant density and plant vigor. Low levels 
of biomass production were caused by low 
plant density. Both plant vigor and bio- 
mass were significantly higher in fall- 
harvested cattails than in spring-har- 
vested cattails. 



Whe 
cattails 
dry matt 
tail sta 
in Minne 
amount o 
wetlands 
of that 
(Bonnewe 
drews 19 
tential 
in catta 
The prod 
related 
biochemi 
in which 
general 
ton Coun 
sediment 
tive cor 



n the dry ma 

in this stu 
er productio 
nds managed 
sota, Wiscon 
f dry matter 

in this stu 
produced in 
11 and Pratt 
80). Thus, 
for increasi 
il wetlands 
uctivity of 
to the nutri 
cal environm 

the cattail 
lower biomas 
ty was relat 

(table 2) . 
relation coe 



tter production of 
dy was compared to the 
n in fertilized cat- 
for biomass production 
sin, and Texas, the 

produced by natural 
dy was less than 50% 
the managed wetlands 

1978, Pratt and An- 
there is a great po- 
ng biomass production 
by proper management, 
cattail wetlands is 
tional status and 
ent of the substrate 
s are growing. In 
s production in Pres- 
ed to the pH of the 

A significant posi- 
fficient (r = 0.73) 



Table 1 — Plant density, plant vigor, and biomass production of cattails in 
wetlands receiving mine drainage. 



Property 



Season 



Monongalia Co. 



Preston Co. 



Plant density 
(plants/m 2 ) 

Plant vigor 
(g dry matter/ 
plant) 



Biomass 
(g dry matter/m 2 ) 



Spring 
Spring 

Fall 

Spring 

Fall 



12.9' 
7.9 - 15.9' 

39.0 
20.1 - 81 .7 



45.5 
25.3 - 91 .5 

411 .9 
211 .9 - 663.0 

554.3 
264.5 - 1180.2 



7-9' , 
0.7 - 19-2^ 

16.5 
11.0 - 24.2 



20.6 
13-2 - 31 .4 

146.2 
5.8 - 345.0 

191 .8 
7.8 - 446.7 



^Mean of seven sites in Monongalia County and five sites in Preston County, 
2 Range. 



361 



was observed between sediment pH and bio- 
mass. Other factors such as low sediment 
concentration of P could also have af- 
fected biomass production. 



taken up by the growing plants. The 
actual amount of metal removal by the 
plant is dependent on the intensity fac- 
tor, i.e. the concentration of metal in 



Table 2 — Properties of sediment in cattail wetlands receiving mine drainage, 



Property 


Season 


pH 


Spring 


Iron^ 


Spring 




Fall 


Manganese^ 


Spring 




Fall 



Monongalia Co, 



Preston Co. 



7.93^ 
7.64 - 8.11^ 

5343 
1992 - 10154 

6418 
2436 - 13,550 

1536 
430 - 2216 

1412 
422 - 2617 



4.67 2 _ 
2.34 - 6.07 3 

3503 
1306 - 7616 

5625 
2464 - 10,850 

280 
50 - 713 

357 
53 - 697 



^0.1 N HC1 extractable iron and manganese (ug/g). 

^Mean of seven sites in Monongalia County and five sites in Preston County. 

^Range. 



Properties of Sediments. 

Wetland sediments have a natural 
tendency to converge to pH 7-0, irrespec- 
tive of the initial soil pH (Sajwan and 
Lindsay 1986). In Preston County wet- 
lands, sediments with pH values lower than 
7.0 indicated the presence of nonequili- 
brium conditions, i.e. these sediments 
represented the wetlands which were con- 
stantly receiving an influx of highly 
acidic water. 

There were significant differences in 
the extractable Mn concentrations between 
the sediments of Monongalia and Preston 
Counties (table 2). Extractable Mn 
concentrations in the Preston County sedi- 
ments were significantly lower than Mn 
concentrations in sediments of Monongalia 
County. The correlation coefficient bet- 
ween the aboveground plant tissue Mn con- 
centrations and concentration of extract- 
able metals in the sediment was 0.52 which 
was nonsignificant. 

No significant differences in the 
amounts of extractable Fe in sediments 
were found between counties or between 
seasons (table 2). No significant corre- 
lation was found between the amount of 
metals extracted from the sediment and the 
concentrations of the metals in the plant 
tissue. This lack of correlation was 
probably due to the nature of the sub- 
strate for plant growth. In wetlands 
receiving acid mine drainage, the extract- 
able metals are indicators of the capacity 
factor for plant nutrition, i.e. the a- 
mount of metal which can be potentially 



the soil solution. In normal plant growth 
media the intensity factor is governed by 
the capacity factor, but in wetlands re- 
ceiving acid mine drainage the intensity 
factor is largely independent of the capa- 
city factor due to differences in water 
chemistry of the drainage and sediment and 
to continuous influx of metals into the 
growth media. This continuous flow of 
mine drainage into the system is responsi- 
ble for the lack of a significant correla- 
tion between extractable metals and con- 
centrations in plant tissue. 

Texture of the sediments in the Mon- 
ongalia County wetlands was sandy loam, 
silt loam, or clay. Texture of wetland 
sediments in Preston Coundy was either 
clay loam or silty clay loam. These tex- 
tures probably represent the textures of 
the surrounding minesoils. 

Concentration and Uptake of 
Manganese and Iron in Cattails. 



Manganese 

The plant tissue analyses for Mn 
(table 3) suggest another reason for the 
yield differential between cattails 
growing in Preston and Monongalia Coun- 
ties. The lower cattail yields in Preston 
County could be due to the high levels of 
Mn in the plant tissue. There was a sig- 
nificant increase in the plant tissue 
concentrations of Mn from spring to fall 
samples for Monongalia County, but the 
seasonal concentration differences were 
not significant for Preston County. This 



362 



lack of increase in Mn concentration with 
growing season in Preston County is pos- 
sibly due to the fact that the concentra- 
tions during the spring season had already 
reached the maximum value for bioconcen- 
tration in the cattail, so there could not 
have been any further concentration in- 
creases . 



at the fall sampling. At the fall 
sampling Fe concentration in Monongalia 
County cattails was 9-2 times greater than 
Fe concentration in Preston County cat- 
tails. These seasonal differences could 
be due to the formation of iron oxide 
coatings that were observed on the roots 
of cattails growing in Preston County. 



Table 3 -- Concentration and uptake of manganese in cattails sampled 
from wetlands receiving mine drainage. 



Plant 
Parts 



Season 



Monongalia Co, 



Preston Co. 



Rhizomes 
(ug/g) 



Spring 



Fall 



Leaves + Stems Spring 
(ug/g) 

Fall 



Leaves + Stems Spring 
(kg/ha) 

Fall 



665 1 
625 - 705 2 

795 
567 - 972 

1202 
702 - 1824 

2202 
586 - 2912 

5.37 
2.64 - 7.92 

11 .47 
3.72 - 23.22 



885 1 _ 
498 - 1278^ 

684 
394 - 1119 

3109 
1879 -4071 

3189 
2653 - 3970 

5.20 
0.11 - 11.60 

6.19 
0.25 - 15.95 



'Mean of seven sites in Monongalia County and five sites in Preston County, 
^Range. 



During the spring season the mean Mn 
concentration in cattails of Preston 
County was 2.5 times greater than the mean 
Mn concentration in cattails of Monongalia 
County (table 3). Even with these large 
differences in concentrations of Mn the 
differences in uptake of Mn at the spring 
sampling between counties were statisti- 
cally nonsignificant. This lack of a 
sizable difference in uptake is related to 
low biomass production in cattail wetlands 
of Preston County. 

Rhizome samples were too few for 
statistical analyses. In general the 
concentrations of Mn in the rhizomes were 
lower than the concentrations in the 
leaves and stems (table 3). In Preston 
County the concentration differences be- 
tween rhizomes and aboveground plant 
portions were four-fold or greater. 

Iron 

Fe concentrations in cattail tissues 
in Preston County were significantly dif- 
ferent at both sampling times from those 
in Monongalia County (table 4). Iron 
concentration in Preston County cattails 
was 3-4 times greater than Fe concentra- 
tion in Monongalia County cattails in the 
spring, but this difference was reversed 



Under reducing conditions, Fe 2+ is thought 
to be oxidized to the less soluble Fe3 + by 
oxidative agents released from the roots, 
thus creating a coating or plaque of an 
insoluble Fe' + compound on the root sur- 
face (Taylor et al. 1984). Iron oxide 
coatings are a survival mechanism for 
cattails growing in a contaminated envi- 
ronment, because the coatings do not allow 
further uptake of Fe (Taylor and Crowder, 
1984, Taylor et al. 1984). Since no addi- 
tional Fe was taken up by the plant, Fe 
concentration decreased as the plant con- 
tinued to grow. Plants growing in Monon- 
galia County may have continued taking up 
Fe during the growing season. The high 
concentrations of Fe found in Preston 
County rhizomes (table 4) are probably due 
to the coatings. 

Seasonal differences in Fe uptake 
were significant for the cattails growing 
in Monongalia County while they were not 
significant for cattails in Preston 
County. In absolute terms, the amounts of 
Fe removed by plants were extremely low, 
i.e. 2.67 and 3-74 kg/ha for the spring 
and fall sampling periods in Monongalia 
County. The amounts taken up in Preston 
County were smaller, i.e. 1.07 kg/ha and 
0.76 kg/ha for the spring and fall sam- 
pling, respectively. 



363 



Table 4 — Concentration and uptake of iron in cattails sampled from wetlands 
receiving mine drainage. 



Plant 
Parts 



Season 



Monongalia Co. 



Preston Co, 



Rhizomes 
(ug/g) 



Spring 



Fall 



Leaves + Stems Spring 
(ug/g) 

Fall 



Leaves + Stems Spring 
(kg/ha) 1 

Fall 



5392^ 
1000 - 10,8183 

595 
109 - 1539 



1919 
409 - 7352 

2.67 
0.23 - 5.47 

3.74 
1 .56 - 5.31 



15.288 2 
9648 - 19,506^ 

11,170 
8036 - 14,650 

2043 
213 - 5369 

208 
88 - 344 

1 .07 
0.04 - 2.80 

0.76 
0.06 - 1 .31 



1 Uptake in kg/ha was calculated by multiplying concentration (ug/g) by biomass, 
2 Rhizome data are means of four sites in each county. All other data are means 
of seven sites in Monongalia County and five sites in Preston Co. 

3Range . 



Removal of Iron from Water by 
Cattail Plants. 

In order to calculate total removal 
of Fe by cattails, biomass of the above- 
ground and belowground parts of the plants 
must be known. Biomass of rhizomes was 
not determined in this study, so it was 
estimated. Biomass of Typha latifolia 
rhizomes varies but it is commonly 50% of 
the total biomass (Personal communication, 
Dr. D.C. Pratt, Botany Department, Univer- 
sity of Minnesota). Therefore, for our 
calculations, we assumed that aboveground 
and belowground biomass were equal. 

To calculate an example of Fe removal 
by cattails in this study, a site in Pres- 
ton County was chosen because it had the 
highest concentration of Fe (19,506 ug/g) 
in the rhizomes. Aboveground and below- 
ground biomass each equal 287.2 g/m 2 . 

To estimate uptake by the rhizomes, 
biomass must be multiplied by the Fe con- 
centration. So, 287.2 g/m 2 X 19,506 ug/g 
= 5,602,123.2 ug/m 2 , or approximately 56 
kg/ha of Fe were removed by the rhizomes 
during the growing season. 

This site had a median flow rate of 
0.5 L/sec, so 15,768,000 L/yr will flow 
into this Preston County wetland. The 
median concentration of Fe entering this 
wetland was 10 mg/L. Therefore 
157,680,000 mg/yr or 157.7 kg of Fe per 
year will be added to this wetland. Since 
the area of this wetland is 61.27 m 2 , 
additional calculations show that approxi- 



mately 25,700 kg/ha of Fe will enter this 
wetland during a year. Since the cattail 
plants are removing only 57.31 kg/ha (56 
kg/ha for the- rhizomes and 1.31 kg/ha for 
the tops), the plants are removing only 
about 0.2% of the total Fe entering the 
wetland. 



Effect of Cattail Wetlands 
on Redox Environment. 



Metals can be 
various mechanisms, 
environment within 
give some informati 
anisms of removal, 
in Preston County, 
230 mV, but the Eh 
other sites ranged 
These Eh values sug 
be a major mechanis 
iron. The pe + pH 
with cattails were 
5), where S0n 2 ~ and 
form insoluble FeSp 
1979). 



removed from water by 

and data on the redox 
a cattail wetland may 
on on possible mech- 
Sediment of one site 
had a positive Eh of 
of sediments at all 
from - 247 to - 18 mV. 
gest that reduction may 
m for the removal of 
values of the sediments 
close to 4.24 (table 
Fe3+ can be reduced to 
at pH 7.0 (Lindsay 



The role of cattails in enhancing the 
reduced environment in wetlands was con- 
firmed when redox potentials were deter- 
mined in wetland sediments with and 
without cattails (table 5). The data 
indicate that the presence of cattails has 
a significant effect on redox environment, 
making it more reducing. Cattails can 
lower the redox potential of the sediments 
enough to precipitate Fe in its sulfide 
form. 



364 



Table 5 — Redox environment in sediment of cattail wetlands receiving mine drainage, 



Property 



Monongalia Co. 



Preston Co, 



PH 



pe + pH 
(with cattails) 

pe + pH 

(without cattails) 



7.95 1 . 
7.64 - 8.11 2 

4.57 
3.93 - 5.36 

10.63 
8.79 - 13.25 



4.63 1 9 
2.34 - 6.07 2 

4.31 
2.62 - 6.20 

9.82 
4.97 - 13-27 



^Mean of seven sites in Monongalia County and five sites in Preston County. 
2 Range. 



CONCLUSIONS 

Amounts of metals removed from mine 
drainage by cattail wetlands depend upon 
cattail biomass and concentration of me- 
tals in plants. In this study biomass 
production appeared to be related to the 
quality of the sediment in the wetland. 
Low pH and high Mn levels were related to 
reduced cattail biomass in some wetlands. 
To increase metal removal from mine 
drainage by plant accumulation, emphasis 
should be placed on increasing dry matter 
production rather than on increased bio- 
concentration. 



Dat 
that pla 
for the 
drainage 
tions sh 
reducing 
cattails 
more red 
duction 
anism by 
Addition 
determin 
metal re 
through 



a from this study a 
nt uptake is not a 
attenuation of meta 
. Eh values and pe 
ow that cattail wet 
environments. The 
in a wetland actua 
ucing environment, 
within the wetland 
which metals like 
al research will be 
e other possible me 
moval from mine dra 
cattail wetlands. 



lso indicate 
major mechanism 
Is in mine 
+ pH calcula- 
lands have 

presence of 
lly promotes a 

Chemical re- 
may be a mech- 
Fe are removed. 

required to 
chanisms of 
inage flowing 



LITERATURE CITED 



In 



Armstrong, W. 1978. Root aeration in 
wetland conditions, pp. 269-291. 
D. D. Hook and R. R. M. Crawford 
(eds.) Plant life in anaerobic envi- 
ronments. Ann Arbor Science. Ann 
Arbor, MI. 

Bonnewell, V. and D. C. Pratt. 1978. 

Effects of nutrients on Tvpha aueus- 
fcjffilia X latifolia productivity and 
morphology. Minnesota Acad, of Sci. 
44:15-20. 

Dahlquist, R. L. and J. W. Knoll. 1978. 
Inductively coupled plasma atomic 
emission spectrometer: analysis of 
biological materials and major trace 
and ultra-trace elements. Applied 
Spectroscopy 39(1 ) : 1- 29. 



Gambrell, R. P. and W. H. Patrick, Jr. 

1978. Chemical and microbiological 
properties of anaerobic soils, pp. 
375-424. In D. D. Hook and R. R. M. 
Crawford (eds.). Plant life in anae- 
robic environments. Ann Arbor 
Science. Ann Arbor, MI. 

Ganje, Y. J. and A. L. Page. 1974. 

Rapid method of dissolution of plant 
tissue for cadmium determination by 
atomic absorption spectrophotometry. 
Atomic Absorption Newsletter 13-131— 
134. 

Girts, M. A. and R. L. P. Kleinmann. 1986. 
Constructed wetlands for treatment 
of acid mine drainage: a preliminary 
review, pp. 165-171. Ill D. H. 
Graves (ed.), 1986 National Symposium 
on Mining, Hydrology, Sedimentology , 
and Reclamation. Univ. of Kentucky, 
Lexington, KY. 

Hutchinson, G. E. 1975. A treatise on 
limnology and botany. Inter- 
science, New York, NY. 

Lindsay, W. L. 1979. Chemical 

equilibria in soils. John Wiley and 
Sons, New York. 449 pp. 

Ponnamperuma , F. N. 1972. The chemistry 
of submerged soils. Adv. in Agron. 
22:29-96. 

Sajwan, K. S. and W. L. Lindsay. 1986. 

Effects of redox on zinc deficiency 
in paddy rice. Soil Sci. Soc. Amer. 
J. 50:1264-1269. 

Sebacher, D. J., R. C. Harris, and K. B. 
Bartlett. 1985. Methane emission 
to the environment through aquatic 
plants. J. Environ. Qual. 14:40-47. 

Soil Survey Staff. 1972. Soil laboratory 
methods and procedures for col- 
lecting soil samples. Soil Survey 
Investigations Report No. 1. U.S. 
Dept. of Agriculture, Soil Conserva- 
tion Service. 63 pp. 



365 



Snyder, C. D. and E. C. Aharrah. 1984. latifolia clones from contaminated 

The influence of the Ty pha community environments. Can. J. Bot. 62:1302- 

on mine drainage. pp. 149-153. In D. 1304. 

H. Graves (ed.), 1984 Symposium on 

Surface Mining, Hydrology, Sedimento- Taylor, G. J., A. A. Crowder, and R. 

logy, and Reclamation. Univ. of Rodden. 1984. Formation and mor- 

Kentucky, Lexington, KY. phology of an iron plaque on the 

roots of Tvpha latifolia grown in 

Taylor, G. J. and A. R. Crowder. 1984. solution culture. Amer. J. Bot. 

Copper and nickel tolerance of Typha 71:666-675. 



366 



WATER AND SOIL PARAMETERS AFFECTING GROWTH OF CATTAILS; PILOT 
STUDIES IN WEST VIRGINIA MINES 1 



David E. Samuel. John C Sencindiver. and Henry W. Rauch' 



Abstract . --Although it has been shown that 
catta i 1 -domi nated wetlands reduce manganese and iron 
content of acid mine drainage, little is known about 
the effects water and soil parameters have on the 
growth of cattails ( Typha 1 ati f o I ia ) . Knowledge of 
such parameters for vegetation growth is essential 
for planned construction of wetlands. Three pilot 
studies were done to: (a) survey water quality where 
cattails grow naturally, (b) compare water and soil 
parameters from sites where cattails were found to 
immediately adjacent spots where no cattails were 
found, and (c) determine the effects of wetland 
vegetation on total iron and manganese concentrations 
and removal in water as compared to an identical site 
with no vegetation. In June 1986 water quality para- 
meters were measured at 26 sites with cattails and 30 
nearby sites without cattails in northern West Virginia 
Water pH was below 2.5 at six cattail sites and below 
2.0 at one site. Only three cattail sites had water 
with over 100 ppm total iron. Comparison of 10 of the 
above-mentioned locations, without any vegetative 
growth, and seven adjacent spots with cattails 
showed that only total iron in water was significantly 
lower in cattail versus non-cattail sites. Data also 
suggest that lower pH and higher aluminum content 
present problems for cattail survival and germination. 
To determine the effects of wetland vegetation on iron 
and manganese removal two watertight boxes (16 ft by 
4 ft by 1 ft) were constructed side-by-side in a cattail 
marsh. Sediment was placed in both boxes and 228 
cattails were transplanted to one on June 17, 1986 
Water entering each box at a flow of 1 gal/ nun had 
total iron concentrations of approximately 20-25 ppm 
and Mn concentrations of 15-20 ppm. Adult cattails 
died, but new sucker growth yielded 261 cattails by 
August 1986. There were 194, 293, and 295 cattails in 
April, May, and June 1987, respectively. From October 
1986 through May 1987, there was an average decrease 
of 3.14 ppm in total iron in cattails and 0.04 ppm in 
the control box. During the same period there was no 
significant decrease in manganese in the cattail box 
(average decrease of 1.9 ppm), or the control. There 
were no changes in pH in either box during the study. 
From the results of these three studies, it appears 
that cattails, and perhaps any wetland vegetation, play 
some role in providing an ecological system which 
functions to remove iron. 



Paper presented at the 1988 Mine 
Drainage and Surface Mine Reclamation Con- 
ference sponsored by the American Society 
for Surface Mining and Reclamation and the 
U.S. Dept. of the Interior (Bureau of Mines 
and Office of Surface Mining Reclamation and 
Enforcement), April 17-22, 1988, Pittsburgh, 
PA. 



367 



INTRODUCTION 

Catta i 1 -domi nated wetlands have been 
shown to reduce manganese (Kleinmann et al. 
1985). and iron (Kleinmann 1985, Brodie 
et al. 1985). Another reason for utiliz- 
ing cattails in manmade wetlands is because 
survival is high. Fulton et al. (1983) 
found that cattail survival after trans- 
planting by various procedures was higher 
(survival was 78%) than 11 other wetland 
species . 

These pilot studies were precipitated 
by discussions with consultants building 
wetlands for industry to improve water 
quality. Little, if anything, is known 
about the various water quality or soil 
chemical properties that may limit the growth 
of wetland plants, including cattails. 
This information is needed so that these 
limitations to growth can be overcome when 
building wetlands to treat acid mine drainage 
While conducting other field studies, we 
noted other sites where cattails were grow- 
ing immediately adjacent to, and in the same 
substrate and seep as, areas where no 
cattails were growing. Therefore, the 
following three pilot studies were done 
to: (a) survey water quality where cattails 
grow naturally; (b)compare water and soil 
parameters from sites where cattails were 
found to immediately adjacent spots where no 
cattails were found; and (c) determine the 
effects of wetland vegetation on total iron 
and manganese concentrations and removal in 
water as compared to an identical site with 
no vegetation. 

This research was funded by a grant 
provided by the U.S. Geological Survey, via 
the West Virginia University Water Research 
Institute. We thank D. McConnell and C. 
Foster for the many hours spent on this 
project. 



MATERIALS AND METHODS 



water 

Morga 

catta 

appea 

the s 

then 

i sola 

to ha 

were 

at ea 

were 

therm 

ture 

condu 



Betwe 

samp 

ntown 

i 1 s g 
red t 
ame s 
i t wa 
ted s 
ve go 
sampl 
ch s i 
made 
omete 
was u 
ct i v i 



en Ju 
les w 
. Si 
rowi n 

o hav 
ubstr 
s sam 
i tes 
od wa 
ed. 
te to 
wi th i 
r fro 
sed i 
ty me 



ne 19 
ere t 
tes w 

9- I 
e the 
ate , 
pled, 
wi th 
ter f 
Water 

the 
n 1 s 
m the 
n sta 
ters . 



and 
aken 
ere s 
fan 

same 
but n 

In 
no ca 
low a 

temp 
neare 
econd 

wate 
ndard 



June 2 
at 56 
electe 
adj ace 

f low 
o catt 
addi ti 
tta i 1 s 
nd a m 
eratur 
st 0.5 

of re 
r. Wa 
izi ng 



6, 19 
s i tes 

d wh i 
nt s i 
of wa 
ails 
on , o 
that 
ud su 

§ c" as 

mo v i n 
ter t 
pH an 



86, 

near 
ch had 
te 

ter and 
growi ng , 
ther 

appeared 
bstrate 

measured 
Readi ngs 
g the 
empera- 
d 



Electrical conductivity was measured 
with an 18-volt Fisher conductivity meter, 
Model No. 152. A Fisher mini pH meter, 
Model No. 640, was used to measure water pH 
Prior to each pH measurement the meter was 



standardized to Ph 4 and pH 7 buffers which 
had been allowed to equilibrate at water 
temperature. The combination electrode 
was cleaned with distilled water between 
each buffer and each water measurement. 
Water flow velocity and depth were measured 
for flow rate calculations to be made at a 
later time. 

Two water samples were collected in 
250-mL nalgene bottles at each sampling 
point. One of these samples was filtered 
through pre-weighted metrical membrane 
filters, pore size 0.45 urn. Both samples 
were treated with 1 mL of 50% nitric acid 
to fix metals. A third unfiltered, 
unacidified water sample was collected in 
1-L plastic cubitainers. 

In the laboratory, unfiltered, 
acidified water samples were used to 
determine concentrations of total igon and 
manganese cations, and of Ca , Mg ' Na 
and K . Filtered, acidified samples were 

used to determine dissolved iron and 
manganese levels. Cation concentrations 
were determined by atomic absorption 
spectro-photometry by ASTM Method D2576 
(ASTM 1976a). 



In 
for the 
sites of 
had dens 
natura 1 1 
either a 
that had 
soil and 
location 
of water 
outl i ned 
from the 
usual ly 
land or 
ment sam 
to a dep 
i nterf ac 
In the 1 
by the d 
Amacher 
absorpti 
was extr 
Howell ( 
spectrop 



the co 
first 

i nter 
e stan 
y, but 
djacen 
no ca 
water 
s with 
ana ly 
above 
same 
i n the 
i n the 
pies w 
th of 
e and 
aborat 
ouble- 
( 1982) 
on spe 
acted 
1978) 
hotome 



urse 
objec 
est w 
d of 

ther 
t to 
ttail 

samp 
i n th 
sis w 
. Se 
poi nt 

mi dd 

nonv 
ere c 
10 cm 
trans 
ory , 
acid 

and 
ctrop 
by th 
and m 
ter. 



of col 
ti ve o 
ere no 
catta i 
e were 
or wit 
s. On 
les we 
ese f o 
ere id 
diment 
s as w 
le of 
egetat 
o 1 lect 

from 
f erred 
cation 
method 
determ 
hotome 
e meth 
easure 



lecti 
f thi 
ted. 
Is gr 

sma 1 
h i n t 

Sept 
re ta 
ur s i 
ent i c 

samp 
ater 
the c 
ed ar 
ed wi 
the s 

to p 
s wer 

of B 
i ned 
try. 
od of 
d on 



ng inform 
s study, 

These si 
owi ng 
1 spots, 
he wetlan 
ember 3, 
ken at 17 
tes. Met 
al to tho 
les were 
samples , 
attai 1 s w 
eas. Sed 
th an aug 
oi 1 -water 
lastic ba 
e extract 
aker and 
by atomic 
Phosphor 
Boltz an 
a UV-VIS 



ation 

four 

tes 



ds, 
1986, 

hods 

se 

taken 

et- 

i- 

er 

gs. 
ed 



ous 
d 



David E. Samuel is Professor of 
Wildlife Biology, John C. Sencindiver is 
Associate Professor of Agronomy, and Henry 
W. Rauch is Professor of Geology. West 
Virginia University, Morgantown, WV. 



To study the effects of wetland 
vegetation on iron and manganese removal, 
two plywood boxes were constructed in the 
middle of a cattail marsh in Greene County, PA 
on June 17, 1986. The marsh is approximately 
4 acres in size and had an average water pH 
of 2.85, total iron of 25 ppm, and total 
manganese of 16 ppm (based on samples 
taken, three locations in three seasons). 

The boxes were 16 ft by 4 ft by 1 ft 
in size, and were imbedded 1 ft deep in the 
marsh, adjacent to each other. They were 
placed so that there was a slight flow of 
water from one end to the other. Inflow 
pipes allowed flow of about 1 gal/min. 
Sediment from the marsh was placed into the 



368 



boxes as substrate. In one box adult Cattail site 

cattail plants were carefully transplanted non-cattail 

at ? the same density found in the marsh (3.5/ A wetland co 

ft ) making a total of 228 cattails. The communicatio 

other box remained devoid of all vegetation. were cattail 

Water samples were taken at the inflow and 2.5 or an ir 

outlfow of each box on June 20 and Forty-two pe 

approximately every month thereafter for a sites in thi 

year (sampling in December 1986 and February the 3.26 mea 

1987 was not possible due to weather con- 2.5 and one 

ditions). Counts of cattail density were cattails wil 

made periodically as well. Sediment samples alone may no 

were taken in July 1987. establishmen 



s had a mean pH of 4.85, while 
sites had a mean pH of 3.26. 
nsultant (B. Pesavento, personal 
n) indicated that only rarely 
s found in water with pH below 
on concentration above 100 ppm. 
rcent ( 1 1/26) of the cattail 
s study had a water pH below 
n, with six sites below pH 
below 2.0. This suggests that 
1 tolerate lower pH, and pH 
t be a limiting factor for the 
t of thi s plant. 



RESULTS AND DISCUSSION 

Where cattails grow. --Water at cattail 
sites had a wider range of pH than water at 
non-cattail sites (fig. 1). 



8.5 



XX 
X 



7.5 



6.5 



5.5 



4.5 



3.5 



2.5 



1.5 



X 
X 

X 
X 
X 

XX 

XX 



4.85 



• •• 

— I 

t 
I 

• •• 

i" 



3.26 



XX 
XX* 



• = Sites without cattails 

* = Sites with cattails 
= Mean pH 



There appears to be a relationship 
between iron concentrations and the presence 
or absence of cattails (fig. 2). More often 
sites with lower concentrations of total 
iron in water had cattails; 57% (15/26) 
of the cattail sites had iron levels below 
25 ppm, as compred with 13% (4/30) of the 
non-cattail sites. In contrast 36.7% (11/ 
30) of the non-cattail sites had iron 
concentrations over 101 ppm (table 1). 

There were only three cattail sites 
with iron concentrations of over 100 ppm, 
with the highest being 144 ppm (table 1). 
On that site only scattered cattails were 
found. A second site with 113 ppm iron had 
only four scattered cattail plants while 
another site with 113 ppm iron had a dense 
stand of cattai 1 s . 



T 
factor 
the si 
by i nc 
h igher 
There 
2.3 (f 
the ei 
2.30 w 
for ei 
3.51 w 
of sit 
highly 
pH sit 
the lo 
noted 
catta i 
s ites 
only t 
ppm or 
for th 
was on 



here m 
s that 
tes wi 
reas i n 

total 
were n 
ig. D 
ght no 
as 233 
ght no 
as 73. 
es wit 

varia 
es wit 
w pH s 
that i 
1 site 
with c 
hree o 

above 
e 12 c 
ly 27. 



ay we 1 1 b 

affect c 

thout cat 

g water p 

iron at 
ine sites 
The av 
n-catta i 1 
.7 ppm wh 
n-cattai 1 
5 ppm (ta 
h pH betw 
ble (tabl 
h no catt 
ites with 
ron level 
s (table 
attai Is h 
f those h 
(table 2 
attai 1 s i 
9 ppm (ta 



e a c 
atta i 
ta i 1 s 
H (ta 
the 1 

with 
erage 

s i te 
i le a 

site 
ble 2 
een 2 
e 2). 
ails 

catt 
s wer 
2). 
ad pH 
ad i r 
). A 
tes w 
ble 2 



ombi na 
1 grow 

were 
ble 2) 
ower p 

pH le 

total 
s wi th 
verage 
s with 
). Ir 
.30 an 
When 
were c 
ails, 
e lowe 
Sevent 

above 
on lev 
verage 
ith pH 
). 



t ion 

th. 

ranke 

, the 

H lev 

ss th 

i ron 

pH b 

tota 

pH a 

on le 

d 3.5 

the 

ompar 

i t wa 

r i n 

een o 

3.00 

els o 

tota 

abov 



of 

When 

d 

re was 

els. 

an 

for 
elow 
1 i ron 
bove 
vel s 
1 were 
low 
ed to 
s 

the 
f 26 
, and 
f 73 
1 i ron 
e 4.02 



No other measured water quality 

parameters (Mn, K, Na, Ca, Mg) showed 

notable trends between sites with and 
out cattai Is . 



any 

with- 



Figure 1.--pH levels of water samples taken 
at 56 sites in northern West Virginia. 



PAIRED SITES 

Water and soil samples were taken on a 
paired basis when possible ( i.e. , one sample 
from a cattail site and another sample from 
an adjacent spot without cattails) to test 
for additional differences associated with 
cattails. Only total iron in water samples 
was significantly different (Wilcoxon's 
sign rank p<0.055) between sites with 



369 



Table 1. — Sites 
concentration 
for pH. 



with and without cattails ranked by Fe 
in water. All values are in ppm units except 



Fe 



747. 


00 


537. 


00 


211. 


30 


169. 


00 


152. 


00 


153. 


70 


133. 


00 


123. 


00 


122. 


00 


116. 


60 


103 


00 


93 


00 


89 


67 


88 


.62 


86 


.00 


85 


.00 


78 


.00 


63 


.50 


63 


.00 


60 


.00 


47 


.63 


39 


.00 


38 


.60 


30 


.80 


27 


.94 


25 


.20 


24 


.40 


23 


.00 


23 


.00 





.38 



Without Cattails 



j>H 



2. 


07 


1 . 


84 


2. 


91 


2. 


36 


3. 


51 


2. 


15 


2. 


14 


5. 


90 


2. 


98 


3. 


57 


2. 


45 


1 


70 


2 


13 


2 


20 


5 


87 


4 


.11 


3 


.54 


2 


.79 


6 


.32 


2 


.33 


2 


.66 


5 


.99 


2 


.54 


2 


.43 


2 


.19 


2 


.47 


2 


.66 


2 


.81 


2 


.65 


7 


.06 



Mn 



34. 


00 


11. 


30 


9. 


30 


2. 


85 


3. 


78 


9. 


12 


3. 


03 


5. 


50 


1 


11 


13 


09 


2 


28 


7 


40 


14 


63 


9 


.90 


5 


.56 


15 


.10 


4 


.22 


7 


.70 


1 


.76 


6 


.30 


6 


.36 


7 


.80 


11 


.10 


4 


.19 


17 


.05 


7 


.40 


15 


.20 


21 


.90 


24 


.90 


2 


.41 



Fe 



144. 


00 


113. 


00 


113. 


00 


91. 


77 


91. 


00 


90 


00 


73 


00 


70 


00 


59 


10 


32 


00 


28 


16 


14 


.90 


12 


.06 


10 


,60 


9 


.80 


8 


.10 


7 


.06 


4 


.98 


4 


.43 


3 


.45 


2 


.14 


1 


.44 


1 


.05 





.39 





.14 



0.02 



With Cattails 



£K 



70 
85 
85 
05 
30 



5.70 
6.78 
2.30 
2.36 
4.02 

2.42 
3.67 
2.54 
3.12 
6.10 

6.41 
2.55 
3.56 
8.25 
3.03 

8.10 
6.10 
7.40 
6.63 

8.25 



7.78 



Mn 

7.30 

3.56 

3.56 

10.08 

15.10 

6.90 

10.30 

3.55 

6.60 

14.50 

16.94 

4.10 

18.92 

51.80 

2.30 

1.46 



15 


80 





42 


5 


40 





24 





.34 





.12 


2 


.41 





.01 



0.04 



catta 
s ites 
(tab! 
diffe 
and M 
both 
(tabl 
by th 
tion 
soi 1 
due t 
secon 
catta 
than 
ppm K 
21 .2 
catta 
to tw 



ils (53.8 
wi thout 
e 3) . Fo 
rences oc 
n (p 0.0 
e lements 
e 3). Th 
e sma 1 1 s 
in sample 
Mn level 
o one s it 
d ha vi ng 
i 1 site s 
12.2 ppm 
, but no 
ppm. Ca 
i 1 si tes 
o samples 



ppm m 
cattai 
r soil 
curred 
62) , w 
presen 
ese di 
amp le 
s. Fo 
found 
e ' s ha 
57.6 p 
oil sa 
Mn. 
other 
level s 
seemed 

of 10 



ean iron 

Is (98.2 

samples 

for Ca 

ith high 

t in sit 

f f erence 

sizes an 

r exampl 

in catta 

ving 184 

pm of Mn 

mples co 

ne catta 

sites we 

in soil 

i nf late 

74 and 7 



) and adj 
ppm mean 
, s igni f i 
(p 0.062 
er levels 
es with c 
s were af 
d great v 
e , the hi 
i 1 s ites 
ppm and 
. No oth 
ntained m 
i 1 site h 
re higher 

samp les 
d as wel 1 
52. 



acent 
i ron ) 

cant 

) 
of 

attai Is 

f ected 

ar ia- 

gh 

was 

a 

er 

ore 

ad 228 
than 

from 

, due 



B. Pesavento (personal communication) 
suggested that aluminum may be toxic to 
cattails with low water pH. Aluminum was 
not measured in water samples, but it was 
in soil samples. The ratio of the average 



pH/Al for 10 soil samples from non-cattail 
sites was .143, while the same ratio from 
7 cattail sites was .240. Thus, it appears 
that lower pH and higher Al of the soil 
may present problems for cattail germina- 
tion or survival. 



soi 1 
level 
to wh 
s ites 
and o 
that 
centr 
Other 
simi 1 
water 
ppm) 
soi 1 
catta 
total 
immed 
a di s 



It is obv 
or water 
s when ca 
en they a 
At sev 
ne withou 
both had 
ations in 

paramete 
ar. One 

pH (2.8- 
in the wa 
pH and hi 
i 1 s thri v 
ly absent 
i ate ly ad 
tinct dem 



10US 

param 
ttail 
re ab 
era 1 
t cat 
h igh 

both 
rs su 
pa i r 
3.0) , 
ter , 
gh so 
ed i n 

-- a 
j a c e n 
arcat 



from t 
eter e 
s are 
sent i 
adj ace 
tails) 
pH and 

soi 1 
ch as 
of sam 

and h 
and si 
i 1 i ro 

one p 
s was 
t. In 
ion be 



able 3 t 
xhibi ts 
present 
n the ad 
nt sites 
, it was 

low i ro 
and wate 
soil Al 
pies sho 
igh iron 
mi larly 
n levels 
lace and 
all vege 

fact, t 
tween th 



hat no one 
low or high 
compared 
jacent 

(one with 

noted 
n con- 
r samples, 
were 
wed low 

( 120-150 
had low 
. Yet 

were 
tation -- 
here was 
e area with 



370 



20 



»15 

B 

55 

o 

§10 

E 

Z 



| ] = Cattail sites (26) 
| = Non-cattail sites (30) 



fe 0-25 ppm 



26-50 ppm 



51-75 ppm 



76-100 ppm 



101+ ppm 



Figure 2. --Iron concentrations in water samples taken at 56 sites 
in northern West Virginia. 



Table 2. — Sites with and without cattails ranked by pH value 
values are in ppm units except for pH. 



All 



Without Cattails 



With Cattails 



pH 



Fe 



Mn 



pH 



Fe 



Mn 



1.70 


93.00 


7.40 


2.05 


91.77 


10.08 


1.84 


537.00 


11.30 


2.30 


91.00 


15.10 


2.07 


747.00 


34.00 


2.30 


70.00 


3.55 


2.13 


89.67 


14.63 


2.36 


59.10 


6.60 


2.14 


133.00 


3.03 


2.42 


28.16 


16.94 


2.15 


153.70 


9.12 


2.54 


12.06 


18.92 


2.19 


27.94 


17.05 


2.55 


7.06 




2.20 


88.62 


9.90 


2.85 


113.00 


3.56 


2.33 


60.00 


6.30 


2.93 


113.00 


4.92 


2.36 


169.00 


2.86 


3.03 


3.45 


5.40 


2.43 


30.80 


4.19 


3.12 


110.60 


51.80 


2.45 


103.00 


2.28 


3.56 


4.98 


15.80 


2.47 


25.20 


7.40 


3.67 


14.90 


4.10 


2.54 


38.60 


11.10 


4.02 


32.00 


14.50 


2.65 


23.00 


24.90 


5.70 


144.00 


7.30 


2.66 


47.63 


6.36 


5.70 


90.00 


6.90 


2.66 


24.40 


15.20 


6.10 


9.80 


2.30 


2.79 


63.50 


7.70 


6.10 


1.44 


0.34 


2.81 


23.00 


21.90 


6.41 


8.10 


1.46 


2.91 


211.30 


9.30 


6.63 


0.39 


2.41 


2.98 


122.00 


1.11 


6.78 


73.00 


10.30 


3.51 


152.00 


3.78 


7.40 


1.05 


0.12 


3.54 


78.00 


4.22 


7.78 


0.02 


0.04 


3.57 


116.60 


13.09 


8.10 


2.14 


0.24 


4.11 


85.00 


15.10 


8.25 


4.43 


0.42 


5.87 


68.00 


5.56 


8.25 


0.14 


0.01 


5.90 


123.00 


5.50 








5.99 


39.00 


7.80 








6.32 


63.00 


1.76 








7.06 


0.38 


2.41 









371 



Table 3. Mean water and soil quality parameters (ppm) for ten sites without cattails 
and seven adjacent sites with cattails. 







Water 


Analysis 




Soil 


Analysis 




Cattail 


sites 


Non- 


-cattail Sites 


Cattail Sites 


Non- 


-cattail sites 


PH 


3.94 






3.87 


4.54 






3.87 


Fe 


53.8 






98.2 


133.6 






93.0 


Mn 


6.3 






6.2 


43.6 






4.4 


K 


8.7 






7.7 


49.0 






8.9 


Na 


18.4 






15.8 


NS 






NS 


Ca 


201.5 






190.1 


439.1 






123.0 


Mg 


76.2 






68.1 


17.5 






10.0 


Al 


NS 






NS 


17.9 






25.7 


Po 


NS 






NS 


5.9 






3.8 


Zn 


NS 






NS 


19.8 






1.6 


Cu 


NS 






NS 


0.5 






0.4 



NS — Not sampled. 



cattails and that with none, within the 
same seep. This was found at several loca- 
tions even though no individual paired 
water or soil parameters were significantly 
different . 

CATTAIL AND CONTROL BOXES 

There has been speculation both in 
academic settings and in the literature, 
about the role of wetland plants in the 
complex process of removing iron and 
manganese from acid mine drainage. Cattail 
plants remove some iron from the environ- 
ment (Snyder and Aharrah 1984), and it is 
known that complex bacterial reactions also 
lead to removal of certain metals from 
water. Some ( e.g. , B. Pesavento, personal 
communication) believe that the species of 
wetland plants is of little consequence in 
improving water quality and that the 
entire wetland ecosystem provides the 
ecological basis for removal of iron and 
manganese. Thus, the plants and the soil 
provide the environment for algae and 
bacteria to remove metals through complex 
oxidation and reduction reactions. 

There were 228 cattails initially 
planted in the 64-ft box on June 17, 1986. 
Three months later, many of the adult plants 
had died, but new sucker growth resulted 
in 261 cattail plants present. Counts were 
also made the first week of April, May, and 
June 1987, yielding 194, 293, and 295 
cattails, respectively. There was an 



increase in plant density from 3.56 cattails 
ft to 4.61 cattails ft in one year. 

Monthly water samples were compared 
for two periods. The first sample was 
taken two days after construction of the 
boxes and placement of substrate and 
cattails. Other samples in this first 
period were taken in July, August, and 
September 1986. In late September, an 
abandoned surface mine located above the 
wetland was opened and water was treated. 
This caused an increase in pH, and a 
decrease in iron and manganese entering 
the inflow of the cattail and control box. 
Thus, a second period was established for 
data analysis beginning in early October 
1986, and running until May 1987. There 
was no significant change in iron or 
manganese during the first period (June- 
September 1986) for either the cattail box 
or the control box (table 4). 

For the second period (October 1986- 
May 1987), there was no significant average 
decrease in the difference (p >0.05) 
between the inflow and outflow for iron in 
the cattail box (decrease of 3.14 ppm iron) 
or the control box (decrease of 0.04 ppm 
i ron) . 

Excluding the first sample, which was 
taken two days after the wetland was con- 
structed, the cattail site showed a reduc- 
tion in iron from inflow to outflow in 
eight of nine samples taken (average monthly 
decrease of 3.46 ppm). Excluding the first 



372 



Table 4. Average inflow and outflow levels (ppm) of iron and manganese 
in a 16-ft by 4-ft system with a flow of 1 gal/min. One system 
had cattails and one had no cattails. In analysis, there were two 
periods*: first, June 20-Sept . 16, 1986 (four samples**); second, 
Oct. 6, 1986-May 21, 1987 (six samples). 



First Period: 



IRON 
Inflow Outflow 



Cattails 


19.53 


17.70 


No Cattails 


20.67 


18.90 


Second Period: 






Cattails 


7.77 


4.63 


No Cattails 


5.42 


5.38 



MANGANESE 
Inflow Outflow 



19.15 
18.57 

12.98 
12.08 



20.23 

20.37 

11.08 
12.32 



* In late September a mine became active above this wetland, and water 
was treated, which raised the pH and lowered the iron and manganese. 
Thus, data are separated. 

**First sample in this period was taken two days after the wetland was 
constructed . 



sample, the control site showed a reduction 
in iron from inflow to outflow in four of 
nine samples (average decrease of 1.43 ppm). 

During the first period, manganese 
increased an average of 1.08 ppm for the 
cattail wetland and 1.80 ppm for the control 
from inflow to outflow. During the second 
period (six samples taken) manganese 
increased at the outflow (average of 0.24 
ppm) in the control but decreased 
(insignificant average of 1.90 ppm at .05 
level) at the outflow of the cattail wet- 
land. The pH range at the inflow was 2.85 
to 3.26 for the first period and 3.65 to 
6.87 for the second, with no significant 
difference in the average monthly change 
for the cattail box or the control at the 
outfal 1 . 

Though preliminary, it appears that 
cattails aid removal of iron. After the 
cattails had been planted and were growing 
for 3 months, there was an average monthly 
decrease in iron from inflow to outflow of 
3.14 ppm. Since only 0.04 ppm was removed 
during this same time in the control, which 
contained sediment only, the plants must 
play some role in providing an ecological 
system which functions to remove iron. 

Soil samples were taken at the inflow 
and the outflow of the cattail and control 
boxes in August 1987 (table 5). Highest 
levels of soil iron were found at the out- 
flow of the cattail box. Levels of soil 
iron were considerably higher in soils 
found in the box with cattails than the 
box with sediment and no cattails. 



CONCLUSIONS 

While its true that iron levels on 
cattail sites might be lower due to the 
plants removing iron, chances are that 
the plants germinate and grow better in 
areas of lower iron and higher pH. 

The paired sites were interesting 
because of the abrupt demarcation within 
the wetland. Bare spots with no vegetation 
were found in the middle of heavy growth of 
cattails. In some of these paired sites 
the soil and water iron and pH levels were 
almost identical yet cattails grew in one 
site and not the other. How the various 
levels of metals in the soil and water 
affect plant germination will require an 
intensive laboratory study. Even then 
interactions between various parameters 
might occur that make results difficult to 
interpret . 

The boxes with cattails and no vegeta- 
tion showed differences in ability to remove 
iron. A study on the relationship to 
cattails, and other vegetation, to bacterial 
production would be most beneficial. These 
pilot studies will not provide answers, 
but they do give us some further direction 
for study. 

LITERATURE CITED 

ASTM. 1976a. Metals in water and waste 
water b atomic absorption spectro- 
photometry. Annual Book of ASTM 
Standards. Water. Part 31. American 
Society for Testing and Materials. 
pp. 350-354. 



373 



Table 5. Soil concentrations (ppm) at the inflow and outflow 
of the cattail box and the control box, sampled August 
4, 1987. Two samples were taken in the wetland near the 
boxes, and averaged. 



Fe 



Mg. 



Al 



£H 



Cattail Box 

Inflow 
Outflow 



71.70 


6.52 


82.50 


3.62 


201.50 


7.92 


68.10 


3.20 



No Cattail Box 

Inflow 52.20 5.26 75.30 3.97 

Outflow 81.40 4.81 21.90 3.68 

Surrounding Wetland 155.00 15.40 53.80 3.87 



Baker, D.E. and M.C. Amacher. 1982. Nickel, 
copper, zinc and calcium, pp. 323- 
336 in A.L. Page et al. (eds), 1982. 
Methods of Soil Analysis. Part 2. 
Chemical and Microbiological Properties, 
Second ed. ASA-SSSA. Mallison, WI. 



Mining, Hydrology, Sedimento logy , and 
Reclamation, University of Kentucky. 
Kexington, KY. 



Brodie, G.A., D.A. Hammer, and D.A. 

Toml j ano v i ch . 1985. Use of manmade 
wetlands to treat acid seepage from 
a coal refuse impoundment. Poster 
Session, in R.P. Brooks, D. E. Samuel, 
J.B. Hill - Teds), Wetlands and water 
management on mined lands: Proceedings 
of a conference at Pennsylvania State 
University, October 23-24, 1985. 



Boltz, D.F. 
metr i c 
2nd Ed 
NY, pp 



and J. A. Howell. 1978. Colori 
determination of nonmetals. 

John Wiley & Sons, New York, 

250-351. 



Fulton, G.W., W.T. Barker, and A. Bjugstad. 
1983. Rooted aquatic plant revegeta- 
tion of strip mine impoundments in 
the northern great plains, pp. 113- 
117 hi M.D. Scott, (ed), Third Biennial 
Plains Aquatic Research Conference, 
Montana State University, August 24-25, 
1983. 



Kleinmann, R.L.P. 1985. Treatment of acid 
mine water by wetlands, pp. 48-52 j_n 
Bureau of Mines, Control of acid mine 
drainage, Information Circular 9027, 
Bureau of Mines. 

Kleinmann, R.L.P. , G.R. Watzlaf, and T.E. 
Ackman. 1985. Treatment of mine 
water to remove manganese. pp. 211- 
217 i_u D.H. Graves, (ed), Proceedings. 
1985 'Symposium on Surface Mining, 
Hydrology, Sed imento logy , and Reclama- 
tion. University of Kentucky, Lexing- 
ton, KY. December 9-13, 1985. 

Snyder, CD. and E.C. Aharrah. 1984. The 
influence of the Typha community on 
mine drainage, pp . 149-153 j_n D.H. 
Graves (ed), 1984 Sympoisum on Surface 



374 



DETERMINING THE CAPACITY FOR METAL RETENTION IN MAN-MADE WETLANDS CONSTRUCTED 

FOR TREATMENT OF COAL MINE DRAINAGE 1 

R. Kelman Wieder 2 



Abstract. — Within the past several years, there 
has been a tremendous increase in the use of man-made 
wetland systems for the treatment of acid coal mine 
drainage. However, quantitative estimates of the 
long-term capacity of a wetland for metal retention 
are lacking. In this paper, an upper limit for Fe 
retention in Sphagnum wetlands is estimated by 
individually considering the biological and chemical 
processes contributing to metal retention in wetland 
ecosystems. Also, different field monitoring schemes 
are discussed in terms of their potential for 
assessing the effectiveness of metal retention in 
man-made wetland systems and their potential for 
extrapolating the long-term capacity for effective 
treatment of mine drainage. Although it has been 
suggested that man-made wetlands may offer a low-cost 
approach to mine drainage treatment, cost/benefit 
analyses cannot be carried out without being able to 
reliably estimate long-term capacity for metal 
retention in a man-made wetland system given a 
particular volume and chemistry of mine drainage 
water. Until long-term capacity for metal retention 
in man-made wetlands can be reliably predicted, the 
environmental and economic potential of wetland 
treatment of coal mine drainage remains difficult to 
assess . 



INTRODUCTION 

One of the most frequently cited 
values of freshwater wetlands is their 
ability to act as nutrient sinks, 
improving the quality of water that flows 
through them (Greeson et al. 1978). This 
characteristic has been exploited in using 
freshwater wetlands for nitrogen and/or 
phosphorus removal from wastewaters 
(Kadlec and Tilton 1979, Godfrey et al. 
1985). Whereas denitrif ication is mainly 
responsible for nitrogen removal from 
wastewaters in wetlands, phosphorus 
retention is accomplished mainly by 
adsorption and precipitation reactions 
with Al, Fe, and Ca in the soil (Nichols 
1983). As such, nitrogen removal does not 



1 Paper presented at the 1988 Mine 
Drainage and Reclamation Conference, April 
17-22, 1988, Pittsburgh, PA. 

2 Associate Professor of Biology, 
Villanova University, Villanova, PA, on a 
leave of absence as a Technical Advisor 
with the U.S. Office of Surface Mining, 
Reclamation, and Enforcement, Eastern 
Field Operations, Pittsburgh, PA. 



necessarily decline with time, but the 
capacity of wetland soils to retain 
phosphorus is limited, and phosphorus 
retention in wetlands decreases over time 
(Richardson 1985) . 

Freshwater wetlands may also act as 
sinks for metals. As a result of several 
field and laboratory studies indicating 
that metals in mine drainage can be 
retained in Sphagnum wetlands (Huntsman et 
al. 1978, Wieder and Lang 1982, 1984, 
Kleinmann et al.' 1983, Burris et al. 1984, 
Gerber et al. 1985, Tarleton et al. 1984, 
Wieder et al. 1985a), there has been an 
increased interest in the potential use of 
man-made wetland systems for the treatment 
of metal-enriched waters, such as those 
often resulting from coal and metal mining 
activity (Girts and Kleinmann 1986) . 
Because metals typically do not have a 
gaseous phase, it is reasonable to assume 
that, like phosphorus, the capacity for 
metal retention in wetland systems is 
finite. However, relatively little effort 
has focused on developing quantitative 
estimates of the ultimate long-term 
capacity for metal retention in man-made 
wetland systems constructed specifically 
for mine drainage treatment. 



375 



Accurate estimation of the ultimate 
capacity for metal retention in man-made 
wetland systems would be useful from both 
practical and regulatory standpoints. 
From a practical point of view, it is 
difficult to carry out cost/benefit 
analyses when the ultimate capacity for 
metal retention in a particular wetland 
receiving a particular volume and 
chemistry of mine drainage is unknown. 
From a regulatory point of view, any 
consideration of bond release based on 
constructed wetland systems for mine 
drainage treatment cannot be objectively 
assessed until reliable predictions of 
long-term effectiveness of wetland 
treatment systems can be made. 



In this 
retention in 
estimated by 
biological an 
contributing 
ecosystems . 
monitoring sc 
of their pote 
effectiveness 
man-made wetl 
potential for 
capacity for 
drainage. 



paper, an upper limit on Fe 
Sphagnum wetlands is 
individually considering the 
d chemical processes 
to Fe retention in wetland 
Also, different field 
hemes are discussed in terms 
ntial for assessing the 

of metal retention within 
and systems and their 

extrapolating the long-term 
effective treatment of mine 



UPPER LIMIT FOR Fe RETENTION 

Five processes are primarily involved 
in Fe retention in Sphagnum wetlands: 
uptake and incorporation of Fe by growing 
Sphagnum mosses, the removal of Fe ions 
from solution by cation exchange, the 
specific adsorption of Fe onto organic 
matter, the formation of insoluble Fe 
oxides, and the formation of insoluble Fe 
sulfides. Based on our understanding of 
the biology and chemistry of each process, 
it is possible to estimate an upper limit 
for Fe retention in Sphagnum wetlands. 

As a plant micronutrient, Fe is taken 
up by growing Sphagnum . The highest 
reported rate of net primary production of 
Sphagnum is 610 g/m 2 /yr (Wieder and Lang 
1983). The highest reported Fe 
concentration in field-collected Sphagnum 
plant tissue is 5.8 mg/g (Rodin and 
Bazilevich 1967) . Multiplying these two 
values gives an estimate of Fe uptake by 
growing Sphagnum of 3.5 g/m 2 /yr. This 
estimate is generous not only because we 
used the highest reported values for both 
growth and tissue Fe concentration, but 
also because Sphagnum growth may be 
inhibited by the high Fe concentrations 
typical of AMD. For example, in a 33 da 
laboratory study, growth in length of S. 
f al lax in solutions containing 100 mg/L Fe 
was reduced by 32.5% relative to control 
plants growing in solutions containing 
mg/L Fe (Kearney 1986) . 

The binding of Fe 2+ to negatively 
charged sites on the peat matrix provides 
another mechanism for Fe retention. We 
have guantified Fe retention through 



cation exchange by placing 5 g of dried 
peat into replicate flasks, and adding 100 
mL of a solution containing either 0, 2, 
4, 6, 8, 12, or 16 meq/L of Fe (added as 
FeS04 ; pH of each solution adjusted to 
4.0). After 16 hr, the mixtures were 
filtered and binding of Fe 2+ with the 
concomitant release of other cations from 
exchange sites was calculated (Fig. 1). 
Using the Langmuir equation, a maximum 
retention of Fe 2+ via cation exchange of 
204 ueq/g (5.7 mg/g) was estimated. 

As with mineral soils (e.g., McLaren 
and Crawford 1973, Miller et al. 1983), 
peat can retain Fe by a chelation-like 
specific binding of the metal cations to 
sites on the organic matter matrix. Using 
0.1 M Na4P20 7 extractions (Wieder and Lang 
1986), the highest organically bound Fe 
concentration that we have ever measured 
in a Sphagnum peat sample is 89.3 mg/g. 
Although little is known about the 
chemical nature of the binding sites on 
Sphagnum peat, it appears that the 
specific binding of Fe to organic matter 
may be a much more important mechanism for 
Fe retention than cation exchange in peat 
exposed to mine drainage. 

Iron can also be retained in peat by 
the formation of insoluble Fe oxides. The 
highest concentrations of amorphous Fe 
oxide (oxalate extractable Fe) and 
crystalline Fe oxide (bicarbonate-citrate- 
dithionite extractable Fe; Wieder and Lang 
1986) that we have ever determined in a 
peat sample are 62.8 and 39.0 mg/g, 
respectively. These peat samples were 
collected from Tub Run Bog, West Virginia, 
a naturally-occurring Sphagnum -dominated 
wetland receiving inputs of acid coal mine 
drainage from an adjacent abandoned coal 
surface mine (Wieder and Lang 1982) . To 
what extent the formation of such Fe 
oxides is abiotic versus biotic is 
presently not known. However, indigenous 
populations of both Fe-oxidizing and Mn- 
oxidizing bacteria have been found in 
field-collected samples of Sphagnum peat 
(Stone 1984) . 

Some studies have suggested that 
bacterial dissimilatory sulfate reduction 
and the formation of Fe sulfides could 
play an important role in the removal of 
Fe from mine drainage (e.g., Tuttle et al. 
1969a, b) . Although sulfate reduction and 
the concomitant formation of Fe sulfides 
does occur in freshwater Sphagnum peat 
(Behr 1985, Behr 1986, Wieder and Lang 
1988) , there is little evidence that Fe 
sulfides accumulate to any significant 
extent in peat exposed to mine drainage 
(Tarleton et al. 1984, Wieder et al. 
1985a, Wieder and Lang 1986) . Presumably, 
the accumulation of Fe sulfides is 
precluded by their reoxidation, although 
at present little is known about the 
process of sulfide oxidation in freshwater 
wetland peat (Wieder and Lang 1988). 
Using the Cr 2+ -reduction technique (Wieder 
et al. 1985b), the highest concentration 
of Fe sulfides (actually H 2 S + S° + FeS2 + 



376 



200 



| 150 



a. 
a. 

§ 100 
a 



Fe ads. 



o 



a. 

ir 
o 

(0 

a 

< 



50 



L 




Mg 2+ des. 



2+ 



EQUILIBRIUM Fe CONCENTRATION (meq/L) 

Figure 1. Adsorption of F e 2+ onto Sphagnum peat by cation exchange and the 
corresponding desorption of Mg 2 +, Ca 2 +, H+, Na+, and K+. 



FeS) that we have ever measured in a 
single peat sample is 0.7 mg/g Fe. 

Assuming that it is possible to 
obtain maximum Fe retention by each of the 
above processes, an upper limit on Fe 
retention in a hypothetical man-made 
Sphagnum wetland, 40 m by 40 m, 30 cm deep 
was calculated (Table 1) . In making these 
calculations, a bulk density for Sphagnum 
peat of 0.1 g/cm was assumed. If such a 
man-made wetland were exposed to mine 
drainage with a flow of 4 L/min and a Fe 
concentration of 100 mg/L (Fe loading rate 
of 576 g/da) , it would take an estimated 
44 years until Fe retention by all 
processes combined would become saturated. 

The analysis in Table 1 reveals that 
incorporation of Fe by growth of Sphagnum 
and retention by cation exchange and 
sulfide formation are relatively minor 
contributors to overall potential Fe 
retention, as compared to the specific 
binding of Fe to organic matter and the 
formation of insoluble Fe oxides. It 
should be noted, however, that 4 L/min 
represents a very low flow. If the flow 
of mine drainage were 40 L/min (still only 
a moderate flow) and the Fe concentration 
200 mg/L, Fe saturation of the wetland 



would occur in only 2.2 years. The time 
estimate obtained in Table 1 also assumes 
that it is possible to obtain the maximum 
Fe retention by each of the processes 
involved. In peat samples that have been 
subjected to mine drainage, either in the 
field or in the laboratory, rarely have we 
observed the maximum saturation of Fe 
retention by all of the processes in Table 
1 (Tarleton et al. 1984, Wieder et al. 
1985a, Wieder and Lang 1986) . To place 
the estimate of the maximum potential Fe 
accumulation in Sphagnum peat in some 
perspective, an Fe concentration of 5,757 
g/m 2 (Table 1) is eguivalent to an Fe 
concentration of 19% of the dry mass! 

The approach used in Table 1 has 
provided an estimate of the ultimate 
capacity for Fe retention in a man-made 
Sphagnum wetland exposed to mine drainage. 
However, data from field situations in 
which wetlands have been constructed 
specifically for mine drainage treatment 
are also needed. In the remainder of this 
paper, we discuss the relative merits and 
drawbacks of different monitoring schemes 
for assessing the effectiveness of wetland 
systems constructed for mine drainage 
treatment. 



377 



Table 1. Maximum Fe retention in a hypothetical man-made Sphagnum wetland, 



Process 



Maximum 
Retention (g/m ) 



Days to Reach 
Saturation 



Growth of Sphagnum 3.5 

Cation exchange 171 

Adsorption onto organic matter 2508 

Formation of amorphous Fe oxides 1884 

Formation of crystalline Fe oxides 1170 

Formation of Fe sulfides 21 

Total 5757 



9.7 


475 


6967 


5233 


3250 


59 



15994 = 44 yr 



MONITORING SCHEME 1 

Periodic Measurement of Inflow and 
Outflow Water Chemistry 

The most commonly employed monitoring 
scheme entails the periodic measurement of 
inflow and outflow water chemistry. 
Obviously, outflow water chemistry is of 
particular interest, since if a wetland 
has been constructed for the treatment of 
coal mine drainage on an active mine, 
Federal and/or State water quality 
criteria must be met for water discharged 
from the mine site. However, outflow 
water chemistry alone provides little if 
any information about the effectiveness of 
the wetland for metal retention. 

Often, both inflow and outflow water 
chemistry are determined, and a "treatment 
efficiency," defined as (Ci - C )/Ci, 
where Ci and C are the inflow and outflow 
concentrations, respectively, for a 
particular metal is calculated (Girts and 
Kleinmann 1986, Girts et al. 1987). 
Although this index of wetland performance 
is intuitively satisfying, site hydrology 
and frequency of sampling are two factors 
that must be taken into consideration when 
interpreting treatment efficiency data. 



The observation th 
concentrations in outfl 
than those in inflow wa 
that the wetland is ret 
is by no means sufficie 
Possible dilution of in 
drainage by unmeasured, 
readily discernible, so 
quality" seepage water 
relatively low metal co 
produce inflated estima 
efficiency within the w 
index of treatment effi 
only in hydrologically 



at metal 

ow water are lower 
ter is suggestive 
aining metals, but 
nt evidence, 
fluent mine 

and perhaps not 
urces of "good 
(i.e. , with 
ncentrations) could 
tes of treatment 
etland. Thus, the 
ciency is valid 
tight wetlands 



where the inflowing mine drainage 
represents the only significant source of 
water and metals and where all of the 
water exits the wetland either at the 
location of the outflow water sample or 
via evapotranspiration. Most man-made 
wetlands are situated in topographically 
low-lying areas, often where seeps are 
common. Moreover, when heavy equipment is 
used during construction of man-made 
wetlands, seepage may be enhanced as a 
result of soil excavation. Thus, it is 
likely that in many instances the 
condition of hydrologic tightness in a 
man-made wetland is not satisfied. 

The calculation of treatment 
efficiency based on mean metal 
concentrations averaged over long periods 
of sampling may mask seasonal patterns 
(i.e., winter versus summer) or more 
short-term patterns such as those that may 
be associated with major rain events. For 
example, it is possible that a wetland may 
be actively retaining metals during 
relatively dry periods (e.g., by the 
formation of metal oxides during periods 
of low water table) , but that accumulated 
metal oxide floes are flushed downstream 
during major rain events (cf. Lang and 
Wieder 1985). Thus, the nature of the 
data base must be taken into consideration 
when interpreting treatment efficiencies. 

Despite the potential problems in 
interpreting data resulting from the 
periodic measurement of inflow and outflow 
chemistry in a man-made wetland, long-term 
changes in treatment efficiency (i.e., a 
progressive decrease in treatment over 
several years) at a particular site may be 
informative, assuming that the hydrologic 
conditions remain relatively constant over 
the entire sampling period. Advantages to 
this monitoring scheme are the minimal 
effort and minimal expense required. 



378 



MONITORING SCHEME 2 

Periodic Measurement of Inflow and Outflow 

Water Chemistry with Measurement of Inflow 

and Outflow Water Fluxes 

The monitoring scheme described above 
can be improved upon substantially by the 
installation of instrumentation to measure 
influent and effluent water fluxes since 
both water and chemical budgets can be 
guantitatively estimated for a man-made 
wetland system. If the estimated quantity 
of water leaving a wetland is less than 
the quantity entering, and if the 
difference is comparable to regional 
estimates of evapotranspiration, then 
dilution of the mine drainage by good 
quality seepage water is probably not 
occurring at the site. 

Once a hydrologic budget is 
calculated and determined to be 
reasonable, then periodic measurements of 
metal concentrations in inflow and outflow 
waters can be used to construct metal 
budgets for the wetland. The accuracy of 
the estimates of net retention/release of 
a particular metal obtained from a metal 
input/output budget is improved as the 
frequency of sampling water for chemical 
analysis increases. Metal budgets can be 
used to evaluate changes in the metal 
retention efficiency of a wetland either 
on a short-term basis (i.e. during and 
following a rain event), seasonally, or on 
an annual basis. If a trend of decreasing 
metal retention is obtained over time, 
extrapolation may provide a quantitative 
way of estimating the long-term capacity 
of a particular wetland for retaining a 
particular metal (cf. Richardson 1985). 

Although the construction of water 
and metal budgets for a wetland provides a 
much better indicator of metal retention 
efficiency than the "treatment efficiency" 
index discussed under Scheme 1, the 
associated cost increase can be 
substantial. Continuous monitoring of 
water flows requires careful and costly 
setup, as well as a considerable time 
commitment involved in subsequent 
maintenance and data reduction. 



MONITORING SCHEME 3 

Periodic Measurement of Inflow and Outflow 
Water Chemistry with Measurement of Inflow 
and Outflow Water Fluxes and Periodic 
Analysis of the Wetland Substrate 



While Scheme 2 repr 
considerable improvement 
with respect to being ab 
metal removal efficiency 
the long-term capacity o 
retain metals, a potenti 
Scheme 2 is that the met 
constructed using occasi 
of metal concentrations 
outflow waters. Thus, i 



esents a 

over Scheme 1 
le to evaluate 

and to project 
f the wetland to 
al problem with 
al budgets are 
onal measurement 
in inflow and 
f the flushing of 



metals from a wetland during major rain 
events is indeed considerable, but rain 
event sampling is not carried out, the 
efflux of metals from the wetland will be 
underestimated in the metal budget 
calculation and wetland efficiency for 
metal retention will be overestimated. 

Periodic determination of metal 
concentration in the wetland substrate can 
be used to verify metal accumulation 
determined from the metal budget 
calculations. Accurate estimation of 
metal accumulation in the substrate 
depends on analyzing a sufficient number 
of substrate samples. In addition, if the 
substrate samples are dried and ground 
prior to chemical analysis, values for 
water content and bulk density of the 
substrate must be obtained in order to 
estimate metal accumulation within the 
entire wetland. Besides providing 
verification for water budget 
calculations, trends indicating increases 
in metal concentrations in the substrate 
over time can also be used to extrapolate 
long-term capacity for metal retention in 
a wetland. 

In comparison to Scheme 2, this 
approach to monitoring a man-made wetland 
involves the additional expense of 
collection and analysis of sediment 
samples. 



DISCUSSION 

It has been suggested that man-made 
wetlands may offer a low-cost approach to 
mine drainage treatment (Wieder and Lang 
1982, Kleinmann et al. 1983, Girts and 
Kleinmann 1986). However, based on the 
analysis in Table 1, it appears that 
wetland treatment of mine drainage has a 
limited potential. Although Table 1 
focuses on Fe retention in Sphagnum 
wetlands, it is likely that, at least 
qualitatively, the findings apply to other 
metals and to other types of wetland 
systems as well. The construction of 
wetlands may be attractive for the 
treatment of seeps with low flow volumes, 
but as the flow of the mine drainage 
increases and/or the metal concentrations 
in the mine drainage increase, the 
effective life-time of a man-made wetland 
will decrease. 

In response to the increased interest 
in using man-made wetlands for mine 
drainage treatment, the U.S. Office of 
Surface Mining, Reclamation and 
Enforcement has adopted the policy that 
man-made wetlands may be used on active 
mine sites as long as a fully operational 
water treatment capability is in place and 
operational downslope from the wetland and 
within the permit area. A second part of 
that policy is that given the present 
state of knowledge, bonds should not be 
released based on the installation of man- 
made wetlands to treat mine drainage. 



379 



From a purely practical point of 
view, if the immediate goal of using a 
man-made wetland for mine drainage 
treatment is to meet Federal and/or State 
water quality criteria, perhaps the single 
most important measurement is water 
quality at the outflow from the wetland. 
Such minimal sampling, however, will 
provide no insight into either the 
efficiency of the wetland for metal 
retention or, and perhaps more 
importantly, the long-term capacity of the 
wetland for metal retention. Until field 
monitoring schemes and/or other approaches 
allow for the quantitative, reliable 
estimation of the long-term capacity for 
metal retention within a wetland given a 
particular volume and chemistry of water, 
the environmental and economic potentials 
of wetland treatment of coal mine drainage 
remain difficult to quantitatively assess. 



ACKNOWLEDGEMENTS 

Technical assistance was provided by 
V.A. Granus, and K.P. Heston, and E.M. 
O'Hara. Portions of this work were sup- 
ported by grants R810082 and R812379 from 
the U.S. Environmental Protection Agency. 



Girts, M.A., R.L.P. Kleinmann, and P.M. 

Erickson. 1987. Performance data on 
T ypha and Sphagnum wetlands con- 
structed to treat coal mine drainage. 
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Drainage Task Force Symposium, 
Morgantown, WV. 

Godfrey, P.J., E.R. Kaynor, S. Pelczarski, 
and J. Benforado (Eds.). 1985. Eco- 
logical Considerations in Wetlands 
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Greeson, P.E., J.R. Clark, and L.E. Clark 
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Values: The State of Our Under- 
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Huntsman, B.E., J.G. Solch, and M.D. 
Porter. 1978. Utilization of 
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coal acid mine drainage abatement. 
Abstracts 91st Ann. Mtg. Geol. Soc. 
Amer., Ottawa, Ontario, Canada. 

Kadlec, R.H., and D.L. Tilton. 1979. The 
use of wetlands as a tertiary treat- 
ment procedure. CRC Crit. Rev. in 
Environ. Control 9:185-212. 



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Tuttle, J.H., P.R. Dugan, and C.I. 

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381 



IMPLICATIONS OF SULFATE-REDUCTION AND PYRITE FORMATION PROCESSES 
FOR WATER QUALITY IN A CONSTRUCTED WETLAND: PRELIMINARY OBSERVATIONS 1 



By Robert S. Hedin, David M. Hyman and Richard W. Hammack^ 



Abstract- - Pre! imi nary observations of water chemistry in 
a recently constructed Typha wetland indicate significant 
improvements in water quality in areas where reduction 
processes are occurring. Where the organic substrate was 
black and loose, sulfate concentrations were 16-59 pet less 
than inflow, iron concentrations were 52-98 pet lower, and pH 
was 2-3 units higher. Ninety-six pet of the iron in these 
areas was in the reduced, ferrous form. Where the organic 
substrate was blanketed with orange iron oxyhydroxides, water 
chemistry was similar to inflow, and dissolved iron wa^ 
predominantly in the oxidized, ferric form. Analyses of the 
organic substrate (mushroom compost) showed an enrichment of 
elemental sulfur in all samples. One substrate sample, 
collected 15 cm beneath the surface, contained both elemen- 
tal sulfur and pyrite. Removal of iron from acid mine 
drainage (AMD) through pyrite formation and storage in anoxic 
sediments may be preferable to removal by oxidation and 
hydrolysis because of pyrite's density, negligible solubility 
in acid water, and placement in the substrate. Researchers 
should consider designs that would emphasize movement of AMD 
through organic-rich, anoxic substrates and thus maximize 
sulfate reduction and pyrite-forming processes. 



INTRODUCTION 

Dissimilatory sulfate reduction and pyrite 
formation are biogeochemical processes that 
conceptually can contribute significantly to the 
effectiveness of constructed wetlands in treating 
acid mine drainage (AMD). In the absence of 
oxygen, bacteria such as Desulfovibrio oxidize 
organic matter using sulfate as an electron 



Ipaper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 1988, 
Pittsburgh, PA. 

^Robert S. Hedin is an Ecologist, post-doc- 
toral fellow; David M. Hyman and Richard W. Hammack 
are Geologists, U.S. Department of the Interior, 
Bureau of Mines, Pittsburgh Research Center, 
Pittsburgh, PA. 



acceptor and subsequently reducing it to hydrogen 
sulfide (see reactions in Table 1). When released 
into the water column, hydrogen sulfide either 
bubbles away as a gas (the common rotten eggs odor 
of many wetlands) or is retained in the wetland as 
sulfides, polysulfides, elemental sulfur, iron 
monosul fides (FeS) and pyrite (FeS^). The 
reduction process produces alkalinity, which 
decreases acidity and raises pH. Formation of 
metal sulfides removes iron from solution. Because 
such changes in water chemistry are the primary 
goals of most AMD treatment systems, a better 
understanding of reducing reactions might result in 
significant improvements in the constructed wetland 
technology. 

In this paper we report preliminary results of 
water and substrate analyses from a recently 
constructed wetland that, in areas, has operating 
sulfate reduction and pyrite formation processes. 
We show the dramatic differences in water chemistry 
that exist in these areas, as compared to inflow 
water or nearby oxic zones. We also show that the 
organic substrate is accumulating reduced sulfur 
compounds and, at depth, pyrite is forming. 



382 



Sulfate Reduction 



SITE DESCRIPTION 



Sulfate reduction is a common anaerobic, 
microbial decomposition process. Its occurrence is 
dependent on the presence of sulfate and suitable 
organic matter in an anoxic, aquatic environment. 
In marine systems sulfate levels are very high, 
and sulfate reduction is only limited by the 
presence of suitable organic matter in anoxic 
sediments (Berner 1984). In most freshwater 
systems, sulfate reduction is limited by very low 
sulfate concentrations and microbially-mediated 
methanogensis is the more important anaerobic, 
decomposition process (Cappenberg 1974, Lovley and 
Klug 1986). Because the formation of iron sulfides 
is primarily dependent on the presence of reduced 
sulfur, freshwater sediments typically have very 
low pyrite levels (Berner and Raiswell 1983). 

Wetlands constructed for the treatment of AMD 
have none of these limitations. During the first 
several years of operations, they contain a huge 
amount of high-quality organic matter. The systems 
typically consist of at least a thousand cubic 
meters of fertile, composted organic matter that is 
spread 15-30 cm deep in a shallow pit, flooded, and 
overlain by 10-20 cm of slowly flowing AMD water. 
Aerobic decomposition of the organic matter at the 
water/substrate interface promotes anoxic condi- 
tions within the flooded organic substrates. The 
acid mine drainage flowing through these wetlands 
is enriched with dissolved sulfate and iron. Thus, 
all the conditions for high rates of sulfate 
reduction and pyrite formation exist. 

The expectation that reduction processes could 
be important in constructed wetlands is not without 
precedent. Tuttle et al . (1969) studied an AMD- 
polluted stream in Ohio that flowed through a 
sawdust dam and was effected by reduction processes 
within the sawdust pile. Water samples collected 
below the dam had sulfate concentrations 60 pet 
lower and iron levels 50 pet less than samples 
collected above the dam. Herlihy and Mills (1985) 
and many of their students have studied a reservoir 
in Virginia that receives AMD from abandoned 
pyrite mines. Summer sulfate reduction rates in 
sediments near the AMD inflow are consistently at 
least a magnitude higher than control areas. Acid 
volatile sulfide concentrations, which are assumed 
to represent iron monosulfides, are always a 
magnitude higher in the AMD impacted sediments than 
in the control sediments. 

Despite these observations, the role of 
reduction processes in constructed wetlands has 
received very little attention. Most research has 
focused on uptake of iron by plants, or 
microbially-mediated metal oxidation processes 
(e.g. Kleinmann et al . 1983, Gerber et al . 1985, 
Wieder and Lang 1986, Kearney and Wieder, elsewhere 
in these proceedings, Webster et al . , elsewhere in 
these proceedings). In the only published studies 
that specifically looked for reduced sulfur 
compounds (and found minor amounts), either the 
samples were collected from the top 20 cm of peat 
in a natural Sphagnum wetland (Wieder and Lang 
1986) or laboratory mesocosms were only 20 cm deep 
(Tarleton et al . 1984). Such shallow depths are 
not conducive to development of the permanent 
anoxic conditions that are necessary to promote 
pyrite accumulation. 



The wetland monitored in this study is located 
in Westmoreland County, Pennsylvania about 16 km 
(10 mi) north of Latrobe. The wetland was con- 
structed in the spring of 1987 to treat drainage 
from recently strip-mined coal spoils (Figure 1). 
It is a cattail -dominated ( Typha ) community growing 
on a substrate of 15 cm (6 in) of crushed lime- 
stone, covered with 45 cm (18 in) of mushroom 
compost^- Water depths over the substrate are 
5-20 cm (2-8 in). Baffles were constructed in each 
cell with bales of hay to promote a serpentine flow 
path. By September, 1987 a dense stand of cattails 
had developed. Seepage drains from the reclaimed 
mine spoils at flows of 40 to 120 L/min (10 to 30 
gals/min). It is collected in a shallow pit, which 
was also filled with substrates and planted with 
cattails, and then flows 30 m (100 ft) over 
limestone rip-rap into the wetland. 




Figure l.--Plan view of the constructed wetland. 



METHODS 

The wetland was examined and sampled in autumn 
of 1987. Chemical measurements were made directly 
in the field and in Bureau of Mines' laboratories 
at the Pittsburgh Research Center. Numerous pH 
determinations were made using a temperature- 
compensated, field pH meter (Orion Model 230) with 
a Ross combination electrode. The electrode and 
meter were calibrated and measurements were 
subsequently made by placing the probe directly 
into wetland water. Dissolved oxygen (DO) was 
measured on October 7 with a portable polarographic 
sensor and field meter (Extech Model 9070). 
Because this DO probe could not be completely 
submerged, measurements reflect DO only to a depth 
of about 5 cm (2 in) beneath the water surface. 



3 Mushroom compost, a waste product of local 
mushroom growers, is produced by aerobically com- 
posting a mixture of spoiled hay, agricultural 
wastes, and animal manure. 



383 



Table 1 . --Important chemical reactions in anaerobic, sulfate-rich environments 

Ref 7 



Reaction description 

1 . Sulfate reduction 

2. hydrogen sulfide 
ionization 

3. formation of 
elemental sulfur 

4. polysulfide 
formation 

5. iron monosulfide 
formation 

6a. pyrite formation 

6b. pyrite formation 

6c. pyrite formation 



Reaction 



S0 4 ^- + 2CH 2 ---> H 2 S + 2HC0 3 " 

H 2 S ---> HS" + H + 
HS" ---> S 2 " + H+ 

HS" + l/20 2 + H + ---> S° + H 2 



HS' + (x-l)S° ---> S x 2 " + H+ 



Fe 2+ + HS" 



■-> FeS + H + 



(1) 
(2) 

( 2 ) 
(2) 



FeS + S° ---> FeS 2 

Fe 2+ + S° + H 2 S ---> FeS 2 + 2H+ 

Fe 2+ + S x 2 " + HS" ---> FeS 2 + S x 2 + H+ (3) 



(4) 

(4) 
(3) 



*CH?0 represents organic matter. 

2 References are: (1) Berner 1984; (2) Chen and Morris 1972; (3) Altschuler 
et al. 1983; (4) Berner 1967. 



Table 2. Changes in water Quality between the main seep and wetland 
inflow on October 27, 1987 

dH Acidity Sul fate Fe Fe^ + Al Hn Na Ca Mg 

main seep 3.04 1000 2125 227 146 49 36 4 188 134 

inflow 3.02 909 2075 203 62 51 38 4 209 139 



Acidity is mg/L of CaC0 3 equivalent; all other values, except pH, 
are mg/L. 



Water samples were collected at various 
locations (Fig. 1) and analyzed in the laboratory 
for pH, acidity, sulfate, Fe tot , Fe 2+ , Mn, Al , Ca, 
Mg, and Na using standard methods (e.g. Watzlaf 
1986). Substrate samples were collected in plastic 
bottles which were filled to capacity, capped, and 
refrigerated. In the laboratory, these samples 
were centrifuged to separate solids from liquid, 
and the supernatant was either analyzed or dis- 
carded. The solids were dried under vacuum and 
stored under argon. Total sulfur was determined 
using a LECO SC-32 Sulfur Determinator. Sulfur 
forms were distinguished using evolved gas analysis 
(EGA) and X-ray photoelectron spectroscopy (XPS) 
(Hammack, elsewhere in these proceedings). X-ray 
diffraction was used to corroborate EGA and XPS 
results. The acid-soluble metal content of solids 
was determined by boiling a weighed sample in aqua 
regia (a mixture of hydrochloric and nitric acids) 
and measuring the metal concentrations of the acid 
extract after filtering it through a Whatman No. 42 
filter. 

RESULTS 

When first examined five months after con- 
struction, the limestone rip-rap in the ditch 
preceding the wetland was coated with iron oxy- 



hydroxides and had little effect on drainage 
chemistry (Tab. 2). On several other occasions we 
measured the pH of water in the pit and at the 
inflow and never found substantial differences. 
The only major change in water chemistry between 
the main seep and the wetland inflow was partial 
oxidation of the ferrous iron (Fe 2+ ). Thus, on all 
sampling dates we have collected wetland inflow 
samples from the end of the riprap (location A in 
Fig. 1). 

Since September, 1987, the wetland has 
displayed a variable appearance. In some areas the 
organic substrate is coated with about 1 cm of an 
orange precipitate that is sometimes hard and 
crusted. In other areas the substrate is black and 
loose. We have visually estimated that black areas 
represent about 20 pet of the wetland area, while 
orange areas represent the rest. Occasionally the 
substrates in black areas are covered with a fine, 
white precipitate. Many of the black areas are 
located in stagnant water behind haybales, while 
orange-colored areas tend to be on the upstream 
sides of haybales and in the central, main-flow 
area. Density of living cattails does not appear 
to be substantially different between the areas. 



384 



The dissolved oxygen (DO) concentrations of 
surface water were lower in black areas than in 
orange areas (Tab. 3). The highest DO value, 10.1 
mg/L or 99% of saturation (water temperature = 13° 
C, saturation = 10.2 mg/L (Truesdale et al . 1955)), 
was found overlying an orange area near the inflow. 
Downstream of the first row of hay bales in the 
first cell, the DO of surface water in an orange 
area was 9.0 mg/L. A black area, only 1 meter 
away, had a surface DO content of 5.0 mg/L. By the 
time water reached the outflow pipe at the end of 
the third cell, DO levels were depressed every- 
where. However, surface water in black areas still 
had 1.2-1.9 mg/L lower DO concentrations than 
nearby orange areas (Tab. 3). Losses of dissolved 
oxygen were likely due to consumption of oxygen by 
aerobic decomposition of organic matter at the 
substrate/water interface, as well as aerobic 
oxidation of ferric iron by chemical and microbial 
processes. The lower values in black areas could 
have resulted from higher oxygen consumption, less 
mixing with oxygenated water, or both. Nowhere 
did we find surface water DO levels low enough to 
be considered anoxic. We suspect, based on sulfur 
and iron analyses (to be discussed), that anoxic 
conditions existed several cm beneath the water/ 
organic substrate interface. 

Table 3. Dissolved Oxygen Content of 
Surface Water in Various 
Environments 

Area description DO (mg/L) 

Seep collection pit 8.7 

First Cell 

Inflow, after riprap 9.2 

Orange area upstream of 

first haybales 10.1 

Orange area downstream of 

first haybales 9.0 

Black area downstream of 

first haybales 5.0 

Third Cell 

Water entering outflow pipe 5.0 

Orange area near outflow pipe... 5.3 

Black area near outflow pipe.... 3.4 

Black area near outflow pipe 4.1 



The orange precipitate that coats much of the 
organic substrate in this wetland was 21% iron 
(Tab. 5) and appears to be an iron oxyhydroxide 
that commonly forms when dissolved iron in AMD is 
oxidized and hydrolyzed. In the first cell, the 
chemistry of surface water flowing through orange 
areas was very similar to inflow water. Numerous 
pH measurements in the first cell (not detailed 
here) consistently found values only several tenths 
of a unit higher than the inflow. The chemistry of 
water collected at the substrate/water interface in 
an orange area at the end of the first cell is 
shown in table 4 (sample B) . Despite having flowed 
through 60 m of wetland, the chemistry was quite 
similar to the inflow (sample A). The only notable 
chemical changes were a decrease in total iron of 
34 mg/L (19 pet of inflow) and an equivalent loss 
of ferrous iron. 

This loss of dissolved ferrous iron, combined 
with the presence of iron oxyhydroxide precipi- 
tates, indicates the predominance of oxidizing 
reactions in these areas. Because abiotic oxida- 
tion of ferrous iron is quite slow at low pH, the 
changes in Fe 2+ in orange areas were presumably 
mediated by iron-oxidizing bacteria. These bacteria 
may also be involved in the removal of iron from 
solution as oxyhydroxides (Gerber et al . 1985). 
The fact that a significant portion of this wetland 
is covered with iron oxyhydroxides indicates the 
historic importance of oxidizing reactions to 
overall iron removal. However, it is also impor- 
tant to note that the current activity of these 
orange areas appears to be limited. Iron is only 
slightly decreased, while all other chemical 
parameters are virtually unchanged. 

Areas in which the substrate surface were 
black had markedly different water chemistry than 
the inflow, or even orange-colored areas located 
less than one meter away. These differences are 
readily seen in table 4 by comparing samples C, D 
and E to A and B. All three samples were collected 
from the substrate/water interface in black areas 
in the first cell (fig. 1). Samples from black 
areas had dissolved iron concentrations that were 
much less than samples at the inflow or nearby 
orange areas. The dissolved iron concentration in 
sample D was only 3 pet of inflow. Over 95 pet of 
the dissolved iron that was observed in black areas 
was in the reduced ferrous form. This contrasted 
markedly with the inflow sample, which was 40 pet 
ferrous, and the nearby orange area sample, which 
was only 27 pet ferrous. 



Table 4. Water Chemistry for Samples Collected October 7, 1987 



SAMPLE 
LOCATION 



CHEMICAL MEASUREMENT 



pH Acidity Sulfate 



AL MEASURE 
Fe Fe^ + 



Al Mn Na Ca Mg 



A Inflow 2.86 872 

B Orange area 2.96 777 

C Black area 4.75 10 

D Black area 5.39 5 

E Black area 4.60 159 
with white precipitate 



2050 181 73 50 36 4 194 133 

2000 147 39 47 36 5 223 139 

1725 30 28 bd 58 5 404 159 

850 3 3 bd 11 6 214 82 

1000 87 84 bd 20 4 171 75 



Acidity is mg/L of CaC03 equivalent; all other values, except pH, are mg/L; 
"bd" is below detection level of instrument (<0.4mg/L). 



385 



Table 5. Acid Soluble Metal Content of Substrate Samples 
Collected at the Water/Substrate Interface on 
October 7, 1987 

SAMPLE METAL CONTENT 1% OF DRY WEIGHT) 

# LOCATION Fe Mn A1 Ca Mg Na 



B Orange Area 

E Black Area 
with white 
precipitate 



21.4 <0.1 
0.2 <0.1 



<0.1 
9.8 



<0.1 
<0.1 



<0.1 <0.1 
<0.1 <0.1 



Percentages do not sum to 100% because of carbon, oxygen, 
hydrogen, sulfur and acid-insoluble metals. 



Dissolved aluminum was decreased to undetec- 
table levels (<0.4 mg/L) in all samples from black 
areas. Sample E represented an attempt to collect 
a fine white precipitate that covered the black 
substrate in several areas. Large amounts of black 
substrate were unavoidably included in the sample. 
Analysis of the acid-soluble constituents of this 
material revealed an aluminum content of almost 10 
pet (tab. 5). Based on the color and this 
analysis, it appears that the white precipitate was 
aluminum hydroxide, which commonly forms when the 
pH of Al -containing water rises above 3.5. Note 
that, unlike the substrate sample taken only one 
meter away in an orange area, iron was a minor 
component of the substrate solids. Despite the 
fact that dissolved iron was very low in the black 
areas, it's loss was not due to precipitation and 
accumulation on the surface of the organic 
substrate. 

Samples in black areas were much less acidic 
than the inflow or nearby orange area samples. 
Additional pH measurements have indicated that the 
surface water in black areas generally has a pH 
between 4.0 and 4.5, while the water within the 
organic substrate often has a pH greater than 6.0. 
This decreased acidity cannot be entirely attrib- 
uted to the neutralizing effect of underlying 
limestone. Two of the three samples had dissolved 
calcium concentrations very similar to the inflow. 
Only sample C had higher calcium levels, but the 
increase over inflow, 5.25 mm/L (millimoles per 
liter), was not enough to explain the 8.6 mm/L 
decrease in acidity (CaC03 equivalents). 



Presumably, 
resulted from sul 
Dissolved sulfate 
and E were lower 
and orange areas 
pet of the inflow 
disturbed, sedime 
wetland released 
bubbles and the c 
sulfide was evide 
sulfide productio 
and acidity is ve 
reduction. 



part of the decreased acidity 
fate reduction (eq.l, tab. 1). 

concentrations of samples C, D, 
than samples from both the inflow 
(Tab. 4). Sample D had only 41 

sulfate concentration. When 
nts in virtually every part of the 
considerable emissions of gas 
haracteristic odor of hydrogen 
nt. This combination hydrogen 
n and decreased dissolved sulfate 
ry strong evidence for sulfate 



Once released into the water column by 
sulfate-reducing bacteria, H?S has several possible 
fates. In acidic water, ionization is low (Eq. 2 
in Tab. 1) and H?S can escape from the wetland as a 
gas. Under alkaline conditions, ionization 
potentially results in significant concentrations 



of both HS" and S2". These dissolved sulfides can 
be oxidized to elemental sulfur by oxygen (Eq. 3) 
or react with ferrous iron to form monosulfides or 
pyrite (Eq. 5 and 6c, respectively). Secondary 
reactions of elemental sulfur can result in 
polysulfides (Eq. 4) and pyrite (Eq. 6a and 6b). 
All of these products precipitate and thus remove 
iron and sulfide from solution. Iron monosulfides 
oxidize very quickly in aerobic environments and 
decompose in acidic solutions to ferrous iron and 
hydrogen sulfide (reverse of eq. 2 and eq. 5). 
Pyrite and elemental sulfur also oxidize to sulfate 
if oxic conditions develop, but neither is soluble 
in the sulfuric acid that gives AMD its acid 
character. 

Currently, judgments about the relative 
importance of each of these sulfur-containing 
compounds in this constructed wetland cannot be 
made. However, it is certain that sulfide reduc- 
tion in excess of that required for iron sulfide 
formation is occurring. Water samples in black 
areas had, on average, 2.5 mm/L (millimoles per 
liter) less iron than inflow water. If all this 
iron was removed by formation of monosulfides and 
pyrite, 2.5-5.0 mm/L of sulfide would have been 
required. Sulfate concentrations in the black 
areas averaged 9.0 mm/L less than inflow samples. 
Thus, at least 4.0 mm/L of reduced sulfur must have 
been either released as gaseous H2S, or retained in 
the wetland in a reduced from not associated with 
iron. 

Except when we disturbed the sediments by 
walking in the wetland, we noticed little odor of 
hydrogen sulfide. This observation may be ex- 
plained by the alkaline conditions that exist in 
the organic substrates and which should cause 
ionization of H?S to the dissolved sulfide forms. 
If losses of sulfide as H2S are not major, then 
retention and accumulation of reduced sulfur in the 
organic substrates must be occurring. We evaluated 
this hypothesis by analyzing samples of the organic 
substrate for sulfur content (LEC0 sulfur analyzer) 
and sulfur form (evolved gas analysis). Mushroom 
compost in its original form has a sulfur content 
of about 1.5 pet, most of which is in the sulfate 
form. All substrate samples collected from this 
wetland have contained more than 2.0% sulfur, and 
we have measured total sulfur contents as high as 
10.2 pet for black particulates filtered from 
wetland water samples. The organic substrate 
filtered from sample D was 3.2 pet sulfur. Evolved 
gas analysis (Fig. 2) of this sample showed that 
most of the sulfur was elemental with subordinate 



386 




250 3S0 

SAMPLE TEMPERATURE. CELSIUS 



0.0017 
0.0016 
0.0019 
0.0014 
0.0013 - 
0.0012 
0.0011 
0.001 
0.0009 
0.0008 
0.0007 
0.0006 - 
0.0003 
0.0004 
0.0003 - 
0.0002 




200 300 400 

SAMPLE TEMPERATURE. CELSIUS 



Figure 2. --Evolved gas analyses of fresh mushroom 
compost (lower line) and substrate samples 
collected from the constructed wetland (upper 
line). The wetland sample was collected from the 
substrate surface at location D. The plotted lines 
indicate evolution of SO? as the samples were 
heated from 50°C to 1000°C (only the 150-500°C 
range is shown). SO2 evolution temperature ranges 
for important sulfur forms are: reduced and 
elemental sulfur, 150-300°C; carbon-bonded sulfur, 
300-350°C; pyritic sulfur, 375-425°C; sulfate, 800- 
900°C. The fresh compost sample had a very large 
peak in the sulfate range that is not shown. 



Figure 3. --Evolved gas analysis of organic 
substrate collected 15-20 cm deep in the 
constructed wetland. The peak at 390°C indicates 
pyrite, while the larger one at 250°C is elemental 
sulfur. 



amounts of organic sulfur that co-evolved with 
aliphatic hydrocarbons. The presence of elemental 
sulfur was confirmed with XPS and X-ray diffrac- 
tion. Elemental sulfur was not present in the 
original compost material (Fig. 2). Pyrite, which 
is indicated by peaks between 380° C and 420° C, 
was not found. This was not unexpected because the 
samples were collected from the surface of sub- 
strates overlain by shallow oxygenated water. 
Authigenic pyrite (pyrite formed in place, not 
transported) is not stable in such an oxidizing 
environment. 

Based on the high iron and reduced sulfur 
concentrations observed in the wetland, pyrite 
would be expected to form in anaerobic areas. We 
tested this prediction by collecting one sample 
from 15-20 cm deep in the organic substrate 
(location F in Figure 1). The sampling area was a 
black pool, and the pH of disturbed water after 
sampling was greater than 6.0. Evolved gas 
analysis of the solid organic material revealed the 
presence of large amounts of elemental sulfur, as 
well as a smaller amount of pyritic sulfur (Figure 
3). The identification of pyrite was confirmed 
with X-ray diffraction analysis. Although we have 
not performed experiments yet to ascertain that 
this pyrite is authigenic (versus being washed in 
from surrounding spoils), we believe that the 
environmental conditions, as well as the pyrite's 
location in the organic substrate argue strongly 
for an authigenic origin. 



DISCUSSION 

Many biogeochemical processes interact to 
cause the amelioration of AMD chemistry as it flows 
through a natural or constructed wetland. Sulfate 
reduction and pyrite formation, though often cited 
as the desirable, ultimate sink for iron removal in 
constructed wetlands, have received little research 
attention. The results in this paper illustrate 
the dramatic chemical changes associated with 
reducing zones in a newly constructed wetland. 

These processes also would appear to have 
important advantages over other mechanisms of AMD 
amelioration. Unlike oxidation processes, which 
produce acidity, sulfate reduction produces 
alkalinity (Eq. 1 in Tab. 1). Sulfate reduction 
and pyrite formation processes are not constrained 
by accumulation limits or toxic effects. Hydrogen 
sulfide is a waste product of dissimilatory 
sulfate-reducing bacteria that is expelled, not 
accumulated. Pyrite formation occurs for solu- 
bility reasons (Howarth 1979) or because of 
chemical transformations in existing inorganic 
sulfur compounds (Berner 1984, Altschuler et al . 
1983). These processes contrast with bioaccumu- 
lation or chelation/adsorption processes. Many 
wetland plants can remove metals from solution, 
however, accumulation is generally limited by 
toxicity factors (Kearney and Wieder, elsewhere in 
these proceedings). Organic matter, such as peat, 
has a well documented tendency to accumulate 
metals, but it is constrained by adsorption 
limitations (Wieder and Lang 1986). 

Pyrite may also be preferable over the iron 
oxyhydroxides that form by oxidation and hydrolysis 
reactions for several reasons. It accumulates 
within the organic substrates and is less likely 
than surface deposits of iron oxyhydroxides to be 
flushed out of the wetland during storm events. As 



387 



long as pyrite stays in a reducing environment, it 
is insoluble in AMD. Iron oxyhydroxides can be 
resolubil ized if the drainage becomes more acidic. 
Pyrite is a quite dense compound whose accumulation 
would not be expected to change flow or mixing 
regimes. Iron oxyhydroxides are not as dense, and 
by accumulating on the surface can fill the wetland 
and deleteriously effect flow and mixing patterns. 

Presently we are unable to estimate the 
importance of reducing reactions as compared to 
oxidizing reactions and plant uptake in this 
wetland. However, if reduction processes are to be 
maximized in constructed wetlands, it is likely 
that designs must be reconsidered. Currently, most 
constructed wetlands have very high surfacervolume 
ratios which maximize aeration of water. The 
result is surface conditions that are conducive for 
oxidation and hydrolysis reactions and, poten- 
tially, the formation of voluminous iron oxy- 
hydroxides on top of the organic substrate. This 
blanket must interfere with mixing of surface and 
subsurface water. Over time, oxidation and 
reduction processes likely become more discon- 
nected, and the possibility exists that the 
importance of reducing reactions will decrease. 
Design features that maximized flow through 
anaerobic environments might be more effective in 
the long term. For example, if inflow pipes were 
positioned at the base of the wetland, water would 
be forced to rise through anoxic organic matter, 
presumably stimulating sulfate reduction and pyrite 
formation. When water reached the surface, it 
would then be "polished" by oxidation and 
hydrolysis reactions. This bottom flow system has 
natural analogies in the New Jersey Pine Barrens 
where sulfate entering Sphagnum wetlands in the 
groundwater is removed by reduction reactions 
before it reaches the surface (Spratt et al . 1987). 



CONCLUSIONS 



Reducing environmen 
sulfur sinks that are n 
wetlands constructed to 
water in oxidized zones 
reducing zones has subs 
tions of sulfate, iron, 
substrates are enriched 
at depth, are accumulat 
conditions do not exist 
a considerable flow of 
surface channels which 
chemistry. Wetland bui 
features that increase 
environments, while res 
factors effect the rate 
pyrite formation in the 



ts represent major iron and 
ot being exploited in most 
treat AMD. Compared to 
of this wetland, water in 
tantially lower concentra- 
and acidity, while organic 
with elemental sulfur and, 
ing pyrite. However, these 
throughout the wetland, and 
water is through oxygenated, 
do not markedly change water 
lders should consider design 
flow through reducing 
earchers must determine what 
s of sulfate reduction and 
se wetlands. 



LITERATURE CITED 

Altschuler, Z. S., M. M. Schnepfe, C. C. Sibler and 
F. 0. Simon. 1983. Sulfur diagenesis in 
Everglades peat and origin of pyrite in coal. 
Science 221:221-227. 

Berner, R. A. 1967. Thermodynamic stability of 
sedimentary iron sulfides. Am. J. Sci. 
265:773-785. 

Berner, R. A. 1984. Sedimentary pyrite formation: 
An update. Geochimica et Cos. Acta. 48:605- 

615. 



Berner, R. A., and R. Raiswell. 1984. C/S method 
for distinguishing freshwater from marine 
sedimentary rocks. Geology 12:365-368. 

Cappenberg, T. E. 1974. Interrelations between 
sul fate-reducing and methane-producing bacteria 
in bottom deposits of a fresh-water lake. I. 
Field observations. Antonie van Leeuwenhoek, 
J. Microbiol. Serol . 40:285-295. 

Chen, K. Y., and J. C. Morris. 1972. Kinetics of 
oxidation of aqueous sulfide by O2. Environ. 
Sci. Tech. 6:529-537. 

Gerber, D. W., J. E. Burns, and R. W. Stone. 1985. 
Removal of dissolved iron and manganese ions 
by a Sphagnum moss system, pp. 365-372 in R. 
Brooks, D. Samuel and J. Hill (eds.), Proc. 
Wetlands and Water Management on Mined Lands, 
23-24 October 1985, Pennsylvania State Univer- 
sity, University Park. 

Herlihy, A. T., and A. L. Mills. 1985. Sulfate 
reduction in freshwater sediments receiving 
acid mine drainage. Appl . Environ. Microbiol. 
49:179-186. 

Kleinmann, R. L. P., T. 0. Tiernan, J. G. Solch, 
and R. L. Harris. 1983. A low-cost low 
maintenance treatment system for acid ine 
drainage using Sphagnum moss and limestone, 
pp. 241-245 in S. B. Carpenter and R. W. 
Devore (eds.), 1983 Sym. Surf. Min. Hydr. Sed. 
Reel., University of Kentucky, Lexington. 

Lovley, D. R., and M. J. Klug. 1986. Model for the 
distribution of sulfate reduction and methano- 
genesis in freshwater sediments. Geochim. 
Ccsmochim. Acta 50:11-18. 

Spratt, H. G., Jr., M. D. Morgan and R. E. Good. 
1987. Sulfate reduction in peat from a 
New Jersey Pinelands cedar swamp. Appl. 
Environ. Microbiol. 53:1406-1411 

Tarleton, A. L., G. E. Lang, and R. K. Wieder. 
1984. Removal of iron from acid mine drainage 
by Sphagnum peat: results from experimental 
laboratory microcosms, pp. 413-420, in D. 
Graves (ed.), 1984 Sym. Surf. Min. Hydrol . 
Sed. Reclam., University of Kentucky, 
Lexington. 

Truesdale, G. A., A. L. Downing, and G. F. Lowden. 
1955. The solubility of oxygen in pure and in 
sea water. J. Appl. Chem 5:53-62. 

Tuttle, J. H., P. R. Dugan, C. B. MacMillan, and 
C. I. Randies. 1969. Microbial dissimilatory 
sulfur cycle in acid mine drainage. 
J. Bacter. 97:594-602. 

Wieder, R. K., and G. E. Lang. 1982. Modification 
of acid mine drainage in a freshwater wetland, 
in B. R. McDonald (ed.), Proceedings of the 
Symposium on Wetlands of the Unglaciated 
Appalachian Region, West Virginia University, 
Morgantown. 

Wieder, R. K., and G. E. Lang. 1986. Fe, Al , Mn, 
and S chemistry of Sphagnum peat in four peat- 
lands with different metal and sulfur input. 
Water, Air, and Soil Pollution 29:309-320. 



388 



AN EVALUATION OF SUBSTRATE TYPES IN CONSTRUCTED 
WETLANDS ACID DRAINAGE TREATMENT SYSTEMS 



Gregory A. Brodie. Donald A. Hammer, and David A. Toml janovich 1 



Abstract--Many wetlands for acid drainage 
treatment have been constructed by the coal and 
utility industries with limited information on 
design and operating criteria. To investigate 
important components (substrates, vegetation, 
microbes) in wetlands treatment systems, the 
Tennessee Valley Authority (TVA) initiated 
experiments at the Acid Drainage Wetlands Research 
Facility. Jackson County. Alabama, in September 
1986. All substrates (6) provided significant 
treatment of dissolved iron, suspended solids, and 
pH. Acid wetland soil was initially more effi- 
cient but differences between substrates became 
insignificant by fall. Only limited manganese 
removal occurred. The pronounced pattern of 
removal efficiency improvement, common to all 
substrate types, suggested that the plant-soil- 
microbial complex important to acid drainage 
treatment developed within all tested substrates 
within one year. Treatment differences between 
substrates were inadeguate to justify added costs 
of deliberately installing a specific substrate in 
operating systems; slightly better performance of 
acid wetland soil supported protecting existing 
wetlands at construction sites. 



INTRODUCTION 

Acid water drainage (pH<6, Fe>4. 
Mn>2), from coal mining, processing, 
transporting, storage, and burning 
lowers water quality, impacts aquatic 



^-Listed alphabetically. Gregory 
A. Brodie is Environmental Engineer. 
Tennessee Valley Authority, Division of 
Fossil and Hydro Power, Chattanooga, TN; 
Donald A. Hammer is Senior Wetlands 
Ecologist, Tennessee Valley Authority , 
Division of Air and Water Resources, 
Norris. TN; David A. Toraljanovich is 
Biologist, Tennessee Valley Authority. 
Division of Air and Water Resources. 
Knoxville. TN. 



biota, and jeopardizes drinking water 
supplies throughout the eastern United 
States. Conventional treatment 
technology consists of grading and 
recontouring to reduce or divert flows 
and addition of alkaline solutions to 
elevate pH levels and chemically 
precipitate metallic ions. Land 
reforming is almost prohibitively 
expensive, and chemical treatment is not 
only expensive but often requires a 
long-term maintenance and operational 
commitment. Constructed wetlands appear 
to offer an inexpensive, self- 
maintaining, long-term solution that may 
be applicable to small or large flow 
acid discharges (Brodie et al. 1986, 
1987). In addition to others in the 
coal and utility industry, the Tennessee 
Valley Authority (TVA) has established 



389 



11 wetland treatment systems at coal 
mines, coal preparation plants, and 
coal-fired plants (Brodie et al 1988). 

Several investigations have 
suggested or demonstrated the effective- 
ness of wetlands in removing acidity, 
sulfate, iron, manganese, and other 
pollutants from acid mine drainage (Holm 

1983, Pesavento 1984. Weider et al. 

1984, Brodie et al. 1987). Some have 
described failures of demonstration 
wetlands to achieve desired results 
(Pesavento 1984, Weider et al. 1985). 

Laboratory studies have explored 
design and operational parameters of 
wetlands for acid mine drainage treat- 
ment (Tarleton et al. 1984, Gerber et 
al. 1985). Little work has been done on 
pollutant removal mechanisms of wetlands 
for mine drainage treatment. Physical/ 
chemical mechanisms may be important, 
but vegetative and microbiological 
mechanisms are thought to be the major 
factors in pollutant removal. Results 
of various treatment systems have ranged 
from poor to excellent, and more impor- 
tantly, few systems have been studied 
over long time periods (Pesavento 1984, 
Weider et al. 1984). Competitive advan- 
tages of man-made wetlands over chemical 
treatment are based not only upon short- 
term economics but upon internal main- 



tenance attributes of wetlands systems 
that suggest long-term independent 
functioning of stabilized systems. 

TVA's Acid Drainage Wetlands 
Research Facility (ADWRF) was estab- 
lished for controlled, pilot-scale 
experiments at an active seep to 
investigate optimal size, type, and 
operational longevity of wetlands 
systems for removing various pollutants 
from acid drainage. Initiated in 
September 1986, the first phase is 
designed to identify most efficient 
substrate type(s). Future studies will 
compare removal efficiencies of different 
wetlands macrophytes and identify 
important microbial populations 
(bacteria, fungi, algae, and protozoa) 
and optimal environmental conditions for 
water quality improvement. 



METHODS 

In August 1986, a series of cells, 
piping, and associated hardware 
comprising the Acid Drainage Wetlands 
Research Facility, was assembled at 
TVA's Fabius Coal Mine Site, Jackson 
County, Alabama. Each of the 20 cells 
consisted of a buried, half-round, 
fiberglass culvert 6.3 m (20 ft) long 
and 1.1 m (42 in) in diameter (Figure 
1). An unlined pond in spoil material 



Individual Cell Cross Section 




Distribution 
Manifold 




jpoo-^ ^-^RQ 

j C ' Ditch/Mixed Plants |0- Sand Bags 

— czsdO 



I Acid Wetland/Bulrush^ t: 



L C5 



Odd-— 



Spoil/Cattail 



Clay/Cattail 



Topsoil/Cattail 



C1 



Normal Wetland/Cattail 



3 



Figure 1. TVA's Acid Drainage Wetlands Research Facility. Jackson County, 
Alabama - Phase 1 Experiment comparing treatment among substrate 
types . 



390 



was planted with marsh/wet meadow 
vegetation. Pond and pea gravel cells 
served as controls. 

Water from a mine seep was pumped 
into a head box and gravity discharged 
into each cell. A polyvinylchloride 
pipe collection system discharged into a 
field scale wetlands treatment system 
and thence to a stream. Appropriate 
valving provided flow control to each 
cell. Seep water had a pH of 5.9 s.u., 
with 37 mg/L iron, and 16 mg/L manganese, 



Substrates 

Clay was obtained from t 
horizon at an undisturbed sit 
facility. Mine spoil was obt 
site. Pea-gravel was purchas 
river gravel operation. Tops 



he B soil 
e near the 
ained on 
ed from a 
oil 

horizon 
ds. Acid 



consisted of soil from the A 
from nearby agricultural fiel 
wetland consisted of substrate from a 
natural wetland below an acid seep 
nearby. Natural wetland consisted of 
substrate from a nearby natural wetland 
without acid drainage. 

Samples of each substrate type 
(except topsoil) were analyzed for 
standard soil composition and chemistry 
parameters (Table 1). After planting, 
ten times the desired level of 0.32 kg 
of 6-12-12 fertilizer was inadvertently 
placed in each cell resulting in an 
eguivalent rate of 4638 kg/ha (4097 lb/ 
acre) . 



Vegetation 

Initial evaluations (1987-1989) 
were designed to test cattail ( Typha 
latif olia ) with bulrush ( Scirpus 
cyperinus ) planted in three cells. 
Planting material was obtained from 
nearby wetlands unimpacted by mine 
drainage. Vegetative material was 
washed to remove soil, cut to standard 
stem length of 30 cm. and weighed before 
planting. 

Fifty-seven cattail, average total 
weight of 6.132 g. were planted on 30 cm 
centers in three rows of 19 plants in 
16 cells on 2 September 1986. Three 
cells were similarly planted to bulrush, 
average total weight of 10,639 g, on 
4 September. Stem densities and heights 
were periodically measured throughout 
the study. 



Monitoring 

Relatively high flow rates to each 
cell (1.0 L/min) were established on 
18 September 1986 then reduced to 0.5 L/ 
min on 20 December 1986 to accelerate 
development of differential treatment 
results. Individual cell flow rates 
were measured, recorded, and adjusted 
daily. 

Weekly (29 September - 5 November 
1986) or twice monthly (after 5 Nov.) 
inflow water samples, obtained from the 
manifolds, and outflow samples from the 



Table 1. Composition of five substrate types tested at the Fabius Acid 
Drainage Wetlands Research Facility, Jackson County, Alabama. 



Substrate Types 





Natural 


Acid 




Mine 


Pea 


Soil Parameters 


Wetland 


Wetland 


Clay 


Spoil 


Gravel 


Organic matter (X) 


0.6 


0.9 


0.5 


0.1 


1.3 


Phosphorus (mg/L) 


5.0 


6.0 


2.0 


3.0 


1.0 


Potassium (mg/L) 


94.0 


98.0 


72.0 


67.0 


1.0 


Magnesium (mg/L) 


114.0 


80.0 


29.0 


37.0 


1.0 


Calcium (mg/L) 


650.0 


290.0 


40.0 


20.0 


4.0 


Sodium (mg/L) 


11.0 


17.0 


9.0 


10.0 


4.0 


pH (mg/L) 


7.0 


4.4 


5.1 


4.8 


5.7 


Nitrate (mg/L) 


4.0 


4.0 


4.0 


6.0 


1.0 


Sulfur (mg/L) 


30.0 


384.0 


41.0 


131.0 


26.0 


Zinc (mg/L) 


2.4 


7.0 


0.9 


2.1 


0.3 


Manganese (mg/L) 


122.0 


45.0 


9.0 


6.0 


1.0 


Iron (mg/L) 


94.0 


138.0 


20.0 


8.0 


1.0 



391 



upper leve 
for pH, te 
conductivi 
(TSS). oxi 
total and 
dissolved 
aluminum, 
tion and a 
followed s 
1979). A 
(ANOVA) wi 
multiple F 
1985) were 



1 of each eel 
mperature, di 
ty, total sus 
dation-reduct 
dissolved iro 
manganese (Mn 
and sulfate, 
nalysis by th 
tandard proce 
two-way analy 
th a Ryan-Ein 
test and pai 
used in data 



1 were 
ssolved 
pended 
ion pot 
n (Fe). 
). ferr 

Sample 
e TVA 1 
dures ( 
sis of 
ot-Gabr 
red t-t 

analys 



analyzed 
oxygen, 

solid6 

ential, 
total and 

ous iron, 
collec- 

aboratory 

US EPA 

variance 

iel-Welsch 

ests (SAS 

is . 



DISCUSSION AND CONCLUSIONS 

Since these results represented 
only the first year of operation of the 
ADWRF, the statistical analyses were 



limited to those parameters most impor- 
tant to operational permit discharge 
limits, i.e.. pH. dissolved Fe and Mn, 
and TSS for the period 14 Jan to 9 Sept 
1987. 

All substrate types significantly 
reduced dissolved Fe and TSS and 
increased pH levels (table 2). Effluent 
concentrations of dissolved Fe ranged 
from 4.8 mg/L (acid wetland/cattail) to 
7.2 mg/L (topsoil/cattail) (table 3). 
Total suspended solids in cell effluents 
ranged from 13.9 mg/L (acid wetland/ 
cattail) to 14.5 mg/L (natural wetland/ 
cattail). Variation in effluent pH 
values was lower, ranging from 6.2 s.u. 
(natural acid wetland/cattail) to 
6.4 s.u. (topsoil/cattail). Effluent 
values for dissolved Mn. ranging from 



Table 2. Comparisons of influent-effluent concentrations of Fe, Mn, TSS, and 
pH between 6 replicated substrates from Jan 14-Sept 9, 1987, at the 
Fabius Acid Drainage Wetlands Research Facility, Jackson County, 
Alabama (paired t-test, 0.05, SAS, 1985). 







Mean Effluent 








Concentration 




Substrate/Vegetation 


N 


(mg/L) 




Prob>'t' 




Influent 


32.3 mg/1 


Dissolved Iron 


Topsoil/Cattail 


57 


7.2 




<0.0001 


Natural Wetland/Cattail 


57 


6.1 




<0.0001 


Clay/Cattail 


57 


5.6 




<0.0001 


Acid Wetland/Bulrush 


57 


5.4 




<0.0001 


Mine Spoil/Cattail 


57 


5.3 




<0.0001 


Acid Wetland/Cattail 


57 


4.8 




<0.0001 




Influent 


14.8 mg/1 


Diss 


olved Manganese 


Natural Wetland/Cattail 


57 


14.5 




0.254 


Clay/Cattail 


57 


14.3 




0.007 


Mine Spoil/Cattail 


57 


14.3 




0.006 


Topsoil/Cattail 


57 


14.2 




0.006 


Acid Wetland/Bulrush 


57 


14.1 




0.038 


Acid Wetland/Cattail 


57 


13.9 




0.076 




Influent 


32.1 mg/1 


Total 


Suspended Solids* 


Topsoil/Cattail 


55 


16.6 




0.008 


Clay/Cattail 


55 


13.6 




0.004 


Acid Wetland/Cattail 


55 


12.2 




0.001 


Mine Spoil/Cattail 


55 


12.1 




0.0006 


Natural Wetland/Cattail 


55 


12.1 




0.002 


Acid Wetland/Bulrush 


55 


11.8 




0.001 




Influent 


5.9 s.u. 




pH** 


Topsoil/Cattail 


56 


6.4 




0.0001 


Acid Wetland/Bulrush 


52 


6.3 




0.0001 


Clay/Cattail 


54 


6.3 




0.0001 


Mine Spoil/Cattail 


53 


6.3 




0.0001 


Natural Wetland/Cattail 


54 


6.2 




0.0001 


Acid Wetland/Cattail 


53 


6.2 




0.0001 



*Two influent (and corresponding effluent) values over 400 mg/1 were excluded 
from the analysis. 
**pH values over 7.5 were excluded from the analysis. 



392 



Table 3. Effect of substrate type on effluent concentrations of dissolved Fe 
and Mn, TSS, and pH from Jan 14-Sept 9, 1987, at the Fabius Acid 
Drainage Wetlands Research Facility, Jackson County, Alabama 
(multiple F test, SAS, 1985). Mean effluent values connected by a 
continuous line are not significantly different. 





Mean Effluent 






Concentration 






(mg/L) 


N 


Dissolved Iron 




5.5 


52 


(rag/1) 




5.2 


54 






5.1 


53 






4.8 


54 






4.7 


54 






4.6 


53 


Dissolved Manganese 




14.3 


53 


(mg/1) 




14.2 


54 






14.1 


54 






14.0 


52 






14.0 


53 






13.9 


54 


Total Suspended Solids 




13.7 


52 


(mg/1) 




11.8 


54 






11.2 


53 






10.7 


54 






10.2 


53 






10.0 


54 


pH 




6.4 


51 


(standard units) 




6.3 


51 






6.3 


49 






6.3 


50 






6.3 


50 






6.2 


49 



Wetland Type 



Topsoil/Cattail 

Clay/Cattail 

Natural Wetland/Cattail 

Mine Spoil/Cattail 

Acid Wetland/Bulrush 

Acid Wetland/Cattail 

Natural Wetland/Cattail 
Mine Spoil/Cattail 
Clay/Cattail 
Topsoil/Cattail 
Acid Wetland/Cattail 
Acid Wetland/Bulrush 

Topsoil/Cattail 

Clay/Cattail 

Acid Wetland/Cattail 

Mine Spoil/Cattail 

Natural Wetland/Cattail 

Acid Wetland/Bulrush 

Topsoil/Cattail 

Clay/Cattail 

Acid Wetland/Bulrush 

Mine Spoil/Cattail 

Natural Wetland/Cattail 

Acid Wetland/Cattail 



13.9 mg/L (acid wetland/cattail) to 
14.5 mg/L (natural wetland/cattail), 
tested significantly different (4 of 
6 substrates) from influent values 
(14.8 mg/L) but absolute removal values 
were very low and statistical signifi- 
cance was inconseguential . 

Seasonal comparisons of effluent 
values revealed that acid wetland/ 
cattail significantly reduced dissolved 
Fe during winter (Jan-Feb) but other 
seasonal differences (Mar-Sept) in Fe or 
in other parameters were unclear. 



Examination of 
these parameters ove 
strong temporal patt 
efficiencies for all 
treatment types that 
overall influent/eff 
2-5). Declines in i 
concentrations, a co 
in pH and extremely 
during the study per 
though possibly rela 
drought conditions. 



the results 
r time expo 
ern in remo 
substrate/ 
was obscur 
luent means 
nfluent Fe 
ncomittant 
variable TS 
iod are une 
ted to regi 



for 
ses a 
val 
plant 
ed by 

(Figures 
and Mn 
increase 
S values 
xplained, 
onal 



Effluent values for TSS. dissolved 
Fe. and Mn suggested that low removal 
efficiencies during the first winter may 
have been due to (1) system initiation 



in late fall and (2) overloading during 
the first 10 weeks of operation. At an 
application rate of 0.5 L/min, effi- 
ciency improved but leveled off until 
warmer weather in early February. 
Removal efficiency gradually improved 
until mid April and rapidly thereafter. 
The pronounced pattern, common to all 
substrate types, suggested that micro- 
bial populations increased dramatically 
with warmer temperatures from mid 
February onward and renewed plant growth 
in April developed system capability to 
remove much of the Fe from the acid 
drainage influent. Results from the pea 
gravel cell showed similar but more 
variable declines in Fe content. 
Figure 2. 



Since the 
statistically s 
substrates, the 
improvement pat 
experimental ce 
biological comp 
process. The p 
complex that is 
guality improve 
without innocul 
tested to date, 
ous nature of d 
organisms in we 
(Gregory and St 



effluent results are not 
ignificant between 

removal efficiency 
tern common to all 
lis suggests a major 
onent in the removal 
lant- substrate-microbe 

important to water 
ment will develop, 
a, in the 6 substrates 

indicating the ubiguit- 
esirable microbial 
tlands treatment systems 
aley 1982). 



393 



50 



40 - 



Sept. 29. 1986 — Sept. 9, 1987 



O 30 - 



Q 20 - 



10 




-l 1 1 1 1 1 1 1 1 — ■ 1 1 1 1 1 — i ; 1 1 1 1- 

9/29 10/14 10/29 11/19 12/17 1/14 2/11 3/12 4/08 5/06 6/03 6/30 7/29 8/26 
FLOW - 1 LITER PER UINUTE I I FLOW = 0.5 LITER PER MINUTE 



MANIFOLD NAT. WETLAND/CATTAIL ACID WETLAND/CATTAIL 

- - TOPSOIL/CATTAIL SPOIL/CATTAIL CLAY/CATTAIL 

PEA GRAVEL/NO PLANTS 

Figure 2. Effect of substrate type on dissolved iron in tests 
conducted at the Acid Drainage Wetlands Research 
Facility. Broken lines represent means of three 
replicated wetland cells, and solid line represents mean 
influent concentrations (manifold). 



Sept. 29, 1986 — Sept. 9, 1987 



D1 

E 



2 




~i — ■ — i — ' — i — ' — r 

9/29 10/14 10/29 11/19 12/17 
FLOW - 1 LITER PER MINUTE 



i — ■ — T 

1/14 2/11 3/12 4/08 
FLOW - 0.5 LITER PER MINUTE 



5/06 6/03 6/30 



r 

7/29 



8/26 



- MANIFOLD 
- - T0PS0IL/CATTAIL 



NAT.WETLAND/CATTAIL 

SPOIL/CATTAIL ' 

PEA GRAVEL/NO PLANTS 



ACID WETLAND/CATTAIL 
CLAY/CATTAIL 



Figure 3. Effect of substrate type on dissolved manganese in tests 
conducted at the Acid Drainage Wetlands Research 
Facility. Broken lines represent means of three 
replicated wetland cells, and solid line represents mean 
influent concentration (manifold). 



394 



1/1 
V) 



60 



50 - 



40 



30 - 



20 



10 - 



Sept. 29. 1986 — Sept. 9, 1987 




-\ — ■ — i — ' — i ■ i • i ■ i ■ i • i — ■ — i — ' — i — ■ — i — > — I — ' — r 

9/29 10/14 10/29 11/19 12/17 1/14 2/11 3/12 4/OB 5/06 6/03 6/30 7/29 8^26 



MANIFOLD NAT.WETLAND/CATTAIL ACID WETLAND/CATTAIL 

- - TOPSOIL/CATTAIL SPOIL/CATTAIL CLAY/CATTAIL 

PEA GRAVEL/NO PLANTS 

Figure 4. Effect of substrate type on total suspended solids in 
tests conducted at the Acid Drainage Wetlands Research 
Facility. Broken lines represent means of three 
replicated wetland cells, and solid line represents mean 
influent concentration (manifold). 



Sept. 29, 1986 — Sept. 9, 1987 



Q. 




9/29 10/14 10/29 11/19 12/17 1/14 2/11 3/12 4/08 5/06 6/03 6/30 7/29 8/26 
aOW - 1 LITER PER UINUTE I I ROW - 0.5 UTER PER UINUTE 



- MANIFOLD NAT.WETLAND/CATTAIL 

-- TOPSOIL/CATTAIL SPOIL/CATTAIL 

' PEA GRAVEL/NO PLANTS 



ACID WETLAND/CATTAIL 
CLAY/CATTAIL. 



Figure 5. Effect of substrate type on pH in tests conducted at the 
Acid Drainage Wetlands Research Facility. Broken lines 
represent means of three replicated wetland cells, and 
solid line represents mean influent concentration 
(manifold) . 



395 



Relatively poor removal of Mn, even 
at the end of the growing season 
compared to much greater removal in many- 
operational systems (Brodie et al. 1988) 
suggested that the plant-soil-microbial 
complex does not substantially alter dis- 
solved Mn compounds until after dissolved 
Fe is relatively low or unavailable for 
microbial metabolism. A lower influent 
application rate resulting in better 
dissolved Fe removal is likely to result 
in improved treatment of dissolved Mn. 

Variations in influent and effluent 
pH seemed closely related although a 
similar gradual improvement over time 
was apparent. In contrast, a strong 
pattern of improved removal of TSS over 
time was evident despite widely varying 
influent values. 

Though trends for pH and dissolved 
Mn are not readily apparent, substantial 
and consistent reductions in dissolved 
Fe and TSS even though influent values 
showed considerable variation, suggest 
that the plant-soil-microbial complex in 
wetlands treatment systems is amenable 
to considerable fluctuation in loading 
rates. 

Interpretation of absolute effluent 
values in terms of wetland treatment 
system efficiencies must incorporate the 
application rates in use during these 



studies. A high application rate (0.5 
L/min. ca 500 ft 2 /gal/min, 0.4 m 2 /mg 
Fe/rain, Brodie et al. 1988) was deliber- 
ately selected to accentuate differences, 
if any, among different substrate types. 
Valve imprecision caused daily average 
flows to vary from 0.34 to 0.43 L/rain 
compared to desired rates of 0.5 L/min. 
Acid and normal wetland substrates 
inadvertantly had higher flow (applica- 
tion) rates. Since these rates are 
substantially greater than recommended 
rates employed at 11 operating treatment 
systems (Brodie et al. 1988), these 
results must not be extrapolated to 
field scale operating systems. 



In addition, catta 
substrate types exhibit 
growth above the substr 
the water column around 
was not present in oper 
treatment systems. Vis 
also revealed a consist 
more vigorous stem and 
the upper portion of ea 
pared to the middle or 



il growing in all 
ed atypical root 
ate and within 

each stem that 
ating wetland 
ual inspection 
ent pattern of 
leaf growth in 
ch cell as corn- 
lower portions. 



Comparison of the average number of 
cattail stems from new shoots present at 
four times during the study revealed 
normal increases throughout the growing 
season but did not show significant 
differences between substrate types 
(Figure 6). Similarly, little differ- 




MAY 7. 1987 



ACID WETLAND CLAY NOR. WETLAND PEA GRAVEL SPOIL 

SUBSTRATE 



Figure 6. Mean number of cattail stems in experimental cells at the 
Acid Drainage Wetlands Research Facility. Jackson County, 
Alabama. One-way ANOVA's on each date showed no 
significant differences among replicated (3 cells) 
substrate types. 



396 



ence was apparent in height of cattail 
grown in different substrate types, 
though topsoil was a significantly- 
better growth medium than pea gravel in 
July and significantly better than spoil 
and pea gravel in October. Analysis of 
phosphorus, total nitrogen and NH 3 / 
NO3 nitrogen showed considerably more 
nitrogen content in the water in the 
upper portion of each cell and a similar 
difference in phosphorus and total 
nitrogen in cattail stems and leaves 
(Table 4) . 

Within-cell differences in cattail 
vigor appeared related to available 
nitrogen and phosphate in seep water and 
relatively low concentrations in sub- 
strate types. Excess fertilizer applied 
at initiation may have flushed out 
leaving influent seep water as the 



principal source for important plant 
nutrients resulting in atypical root 
growth in the water column. Since all 
cells exhibited atypical root growth and 
differences in cattail vigor were 
greater within each cell than between 
cells, these variables were unlikely to 
have differentially influenced the 
comparison of removal efficiencies in 
substrate types. 

In summary, our results suggested 
(1) that substrate type is relatively 
unimportant to removal treatment 
efficiency since the desired plant- 
substrate-microbe complex became 
established in each type; (2) that 
microbial innocula were unnecessary; and 
(3) that vegetation may substantially 
improve treatment efficiency. From a 
practical standpoint, substantial 



Table 4. Nutrients in water and sediment (October 10, 1987) and cattails 

(October 16, 1987) in three experimental wetland cells at the Acid 
Drainage Wetlands Research Facility, Jackson County, Alabama. 



Wetland 


Location 


Water 


Sediment 


Cattail 


Cell 


in Cell 


(mg/L) 


(mg/kg) 


mg/kg dry wt) 






Total Phosphorus 






A5 


upper 


0.01 


170 


420 


A5 


middle 


0.02 


280 


320 


A5 


lower 


0.03 


180 


260 


B6 


upper 


0.01 


240 


520 


B6 


middle 


0.05 


400 


320 


B6 


lower 


0.03 


120 


200 


C5 


upper 


<0.01 


350 


960 


C5 


middle 


0.01 


240 


400 


C5 


lower 


0.02 
Total Kjeldahl Nitre 


170 

gen 


320 


A5 


upper 


0.20 


330 


4900 


A5 


middle 


<0.02 


310 


3740 


A5 


lower 


<0.02 


310 


2720 


B6 


upper 


0.20 


430 


5810 


B6 


middle 


<0.02 


430 


1980 


B6 


lower 


<0.02 


310 


2780 


C5 


upper 


0.16 


340 


7880 


C5 


middle 


<0.02 


360 


3380 


C5 


lower 


<0.02 
Ammonia Nitrogen/Nitrate 


250 
Nitrogen 


4870 



A5 
A5 

A5 



upper 

middle 

lower 



0.51/0.02 
<0. 01/0. 01 
<0.01/<0.01 



B6 
B6 
B6 



upper 

middle 

lower 



0.51/0.02 
0.01/0.02 
0.01/0.01 



C5 
C5 
C5 



upper 

middle 

lower 



0.50/0.03 
<0. 01/0. 02 
<0. 01/0. 01 



397 



differences must exist between substrate 
removal efficiencies to justify the 
considerable construction costs entailed 
in deliberately installing a specific 
substrate in field scale operating 
systems . 



The functional va 
vegetation previously 
operating system resul 
by analysis of these d 
strates with emergent 
higher removal efficie 
strates lacking vegeta 
understanding of wetla 
anticipated from the r 
comparisons of removal 
different wetlands veg 
Acid Drainage Wetlands 



lues of wetlands 
identified from 
ts were supported 
ata. i.e., sub- 
vegetation had 
ncies than sub- 
tion. A better 
nds vegetation is 
esults of future 

efficiencies of 
etation at the 

Research Facility. 



LITERATURE CITED 

Brodie. G. A., D. A. Hammer, and D. A. 
Toml janovich. 1986. Man-made 
Wetlands for Acid Drainage 
Control. Proc. 5th Ann. Natl. 
Abandoned Mine Lands Conf. 10-15 
Aug. 1986. Billings, Montana. 

Brodie, G. A., D. A. Hammer, and D. A. 
Toml janovich. 1987. Treatment of 
Acid Drainage from Coal-Facilities 
with Man-Made Wetlands. I_n Reddy, 
K. R. and W. H. Smith (eds). 
Aquatic Plants for Water Treatment 
and Resource Recovery. Magnolia 
Publ., Orlando. 1032 p. 

Brodie, G. A.. D. A. Hammer, and D. A. 
Toml janovich. 1988. Constructed 
Wetlands for Acid Drainage Control 
by the Tennessee Valley Authority. 
Proc. 1988 Mine Drainage and 
Reclamation Conference, April 17-22 
Pittsburg. PA. 

Federal Water Pollution Control 

Administration. 1969. Stream 
Pollution by Coal Mine Drainage in 
Appalachia. U.S. Dept. of 
Interior. Wash. D.c. 

Gerber, D. W. . J. E. Burris, and B. W. 

Stone. 1985. Removal of Dissolved 
Iron and Manganese Ions by a 
Sphagnum Moss System. I_n: R. P. 
Brooks et al. (eds.). Wetlands and 
Water Management on Mined Lands, 
Proc. of a Conf., October 1985, 
Pennsylvania State University. 



Gregory. E., and J. T. Staley. 1982. 

Widespread Distribution of Ability 
to Oxidize Manganese Among 
Freshwater Bacteria. Applied and 
Environmental Microbiology, 44:2. 

Holm. J. D. 1983. Passive Mine 

Drainage Treatment: Selected case 
studies. I_n: Medin, A. and M. 
Anderson (eds.), Proc. of the ASCE 
Specialty Conference, 1983, 
National Conference on 
Environmental Engineering, Boulder, 
CO. 

Pesavento, B. G. 1984. Factors to be 
Considered when Constructing 
Wetlands for Utilization as Biomass 
Filters to Remove Minerals from 
Solution. In.: J. E. Burris, 
Treatment of Mine Drainage by 
Wetlands, Contribution No. 264 of 
the Department of Biology, 
Pennsylvania State University. 

Statistical Analysis System. 1985. SAS 
Users Guide: Statistics, Version 5 
Ed. Cary. NC-SAS Inst., Inc. 956 p. 

Tarleton. A. L.. G. E. Lang, and R. K. 
Weider. 1984. Removal of Iron 
from Acid Mine Drainage by Sphagnum 
Peat: Results of Experimental 
Laboratory Microcosms. Proc. 
National Symp. Surface Mining, 
Hydrology, Sedimentology, and 
Reclamation, University of Kentucky. 

US EPA. 1979. Methods for Chemical 

Treatment of Water and Wastewater. 
EPA 600/4-79-020. 

Weider. R. K.. G. E. Lang, and A. E. 
Whitehouse. 1984. The Use of 
Freshwater Wetlands to Treat Acid 
Mine Drainage. In.: Burris J. E. 
(ed.). Treatment of Mine Drainage 
by Wetlands. Contribution No. 264 
of the Department of Biology. 
Pennsylvania State University. 

Weider. R. K. . G. E. Lang, and A. E. 

Whitehouse. 1985. Metal Removal 
in a Sphagnum-Dominated Wetlands. 
In : Brooks. R. P.. et al. (eds.). 
Wetlands and Water Management on 
Mined Lands Process of a 
Conference, October 1985. 
Pennsylvania State University. 



398 



ISOLATION AND CULTURE OF A MANGANESE-OXIDIZING BACTERIUM FROM A MAN-MADE CATTAIL WETLAND 

Vail, W. J. (1), Wilson, S. (2), and Riley, R. K. (3). ((1) Professor and Chairman, (2) Student, and 
(3) Professor, Department of Biology, Frostburg State University, Frostburg, MD). A manganese-oxidizing 
bacterium was isolated and cultured in association with a fungus from a cattail wetland. Sediment and 
water samples were collected from a man-made cattail wetland in Somerset County, PA during the summer of 
1987. Winogradsky columns were prepared, and the columns were sampled and transferred to tubes containing 
a medium composed of manganese sulfate and yeast extract at pH 4. After 3 days incubation a fungus 
mycelium appeared, and after 3 additional days incubation the mycelium began to turn dark. Microscopic 
examination showed many regions associated with the hyphae containing dark brown crystalline deposits 
which are assumed to be manganese dioxide. The fungus grown on the control medium, which lacked manganese 
sulfate, produced a cream-colored mycelium, but did not exhibit these brown deposits. Using standard 
microbiological techniques, a fungus-bacterium association was isolated, and observed and confirmed with 
SEM. The oxidization of Mn occurred at pH 4 and above. When this association is inoculated into a medium 
of manganese sulfate and yeast extract, there is a reduction of [Mn] in the supernatant fluid of between 
18 to 26 percent over a 12-day period. The bacterium was tentatively identified as Metallogenium , and 
attempts are underway to grow both the fungus and bacterium in separate cultures. 

Additional Key Words: Metallogenium . 

SCREENING OF MOSSES AND ALGAE IN GREENHOUSE EXPERIMENTS FOR THEIR ABILITY TO REMOVE IRON FROM WATER. 

Webster, H. J. (1), Stark, L. R. (2), and Stevens, S. E., Jr. (3). ((1) Assistant Professor, Biology 
Department, Dubois Campus, The Pennsylvania State University, DuBois, PA 15801, (2) Research Associate, 
Biology Department, The Pennsylvania State University, University Park, PA 16801 and (3) Professor, 
Molecular and Cell Biology, The Pennsylvania State University, University Park, PA 16802). Greenhouse 
experiments were conducted to screen the potential of five moss species, one algal species, and sawdust 
for their effectiveness in lowering the concentration of iron in acidified water. The initial iron 
concentrations of the reservoir water were 37—45 mg/L, using FeSOij as the source of iron. The pH for each 
species was adjusted to the pH of the water at the collection sites as follows: Sphagnum recurvum (5.2), 
S. f imbriatum (5.5), Drepanocladus fluitans (5.2), Pohlia nutans (2.8), Aulacomnium palustre (5.5), and 
Ulothrix subtilissima (3.1), while the red oak ( Quercus borealis ) sawdust was screened at a pH of 5.5. 
Water flow rates were 60 mL/min through each system. The percent change in iron concentrations of the 
water varied with species, although direct comparisons among species should be made with caution. After 
13 to 14 hours, the dissolved - iron in outlet water had been lowered relative to inlet water as follows: 
Spagnum recurvum (92%), S. f imbriatum (78$), Drepanocladus (86%), Pohlia (50?), Aulacomnium (98?), 
Ulothrix (27%), and sawdust (49%). Iron-oxidizing microbes were associated with the plant species 
although no correlation was found between bacterial counts and the lowering of iron concentrations. The 
iron content of plant tissue was> greatest near the inlet of each lane, and brown tissues had more iron 
than green tissues. These results suggest that certain mosses and algae may be effective in treating acid 
mine drainage. 

Additional Key Words: wetlands, Sphanum recurvum , S. f imbriatum , Drepanocladus fluitans , Pohlia nutans , 
Aulacomnium palustre, Ulothrix subtilissima. 



ECOLOGICAL ENGINEERING AND BIOLOGICAL POLISHING: ITS APPLICATION IN CLOSE-OUT OF ACID GENERATING WASTE 
MANAGEMENT AREAS 

Kalin, M. (Ecologist, Boojum Research Limited, Toronto, Canada). Methods for establishing cattail stands 
on pyritic wastes are being developed as part of Ecological Engineering. Cattails reduce infiltration of 
precipitation and provide organic matter. Hand-transplanted, mature Typha latifolia L. developed into 
stands with significantly higher growth potentials when transplanted in groups comprised of >3 plants. 
Mechanical transplanting was more suitable in physically exposed locations, but was limited by machinery 
access. Although hydroponic transplanting failed in waters with high Zn and Cu concentrations (250 and 
35 mg/L), growth was achieved in waters with 3 and 0.5 mg/L Zn and Cu. Effective metal removal by aquatic 
biota (biological polishing) from acidic waste water is a function of the annual biomass production and 
their adsorption/uptake characteristics. Biomass production of Drepanocladus fluitans (C. Muell.) Roth, 
transplanted in mesh bags into an acidified lake (pH 4.0) was seasonally dependent (maximum, 100 g [dry 
weight]/m 2 moss; July, August), but not affected by location within the lake. Its polishing capacity 
ranges between 0.05 to 0.1, and 0.01 to 0.03 g/m 2 Zn and Cu. The biomass production of a gelatinous algal 
community, primarily Muogeotia Agardh spp., colonizing a substrate of submerged woody debris in the same 
lake, was dependent on location, but not on season. Biomass production in the first year of substrate 
introduction ranged from 9 to 42 g [dry weight]/100 g branch, while the polishing capacity of this 
periphytic algae-branch complex after 6 months of growth was 0.1 and 0.03 g /100 g branch Zn and Cu. 
Effective methods curtailing environmental degradation due to acid-generating base metal mining wastes can 
be developed through combining different polishing agents. 

Additional Key Words: base metal mining, mine waste water, copper contamination, zinc contamination, 
Typha latifolia, Drepanocladus fluitans, Muogeotia spp. 



399 



WETLANDS AND MINE DRAINAGE - AN ECOTECHNOLOGY APPROACH 

Mitsch, W. J. (1), Fennessy, M. S. (2), Cardamone M. A. (3), and Palmieri, D. (H). ((1) Professor, School 
of Natural Resources, The Ohio State University, Columbus, OH 43221 , (2) Graduate Assistant, School of 
Natural Resources, The Ohio State University, Columbus, OH 43221, (3) Assistant Director, Wetlands 
Research Inc., 53 West Jackson, Chicago, IL 60604, and (4) Graduate Assistant, School of Natural 
Resources, The Ohio State University, Columbus, OH 43221. Ecotechnology and ecological engineering mean 
the purposeful matching and design of solar-powered ecosystems for benefit of both humans and nature. The 
control of acid mine drainage offers significant opportunities to explore the principles of ecotechnology. 
A general framework of the use of natural ecosystems for the "treatment" of water is presented, based on 
several years of wetland research in Kentucky and Ohio, followed by specifics of a case study in eastern 
Ohio. The case study involves a constructed, 0.25 _ ha wetland, organized into three cells of Typha 
latifolia , which is being used as an alternative control for acid mine drainage. Changing water quality 
and vegetation characteristics have provided a measure of the effectiveness of the wetland in removing 
dissolved iron from the mine effluent. Total reduction of iron from the mine water has averaged 57% 
within the system, with the largest proportion (21 .3%) of the iron removal occurring in the third and 
final wetland cell, where vegetation density is greatest. The rate of iron removal is lower in the first 
two wetland cells, where vegetation density is also lower. Distribution of vegetation appears to be 
independent of distance from the mine effluent; it is controlled by the depth of standing water in the 
wetland. Growth (height and number of leaf blades) does not appear to be significantly affected by the 
mine water. If ecotechnology is to be successful in helping to control acid mine drainage, then design 
parameters and empirical models must be developed for the use of wetland ecosystems. 

Additional Key Words: iron removal, Typha latifolia. 



PEAT BLANKET TO LOCK ACID MINE SPOILS IN A SELF-SUSTAINING ECOSYSTEM 

Brown, A. (1), Mathur, S. P. (2), and Kushner, D. J. (3). ((1) Graduate Student, Department of Biology, 
University of Ottawa, Ont., Canada, (2) Research Associate, Land Resources Research Centre, Agriculture 
Canada, Ottawa, and (3) Professor, Department of Biology, University of Ottawa, Ont., Canada). Acid 
leaching from pyritic mine tailings is due to microbial oxidation of iron sulphides which provides the 
energy source for the organisms. This conversion of sulphur to sulphate, and ferrous to ferric iron, is 
very inefficient and so requires a good supply of oxygen. Exclusion of oxygen from the tailings thus 
prevents acid leaching. It is proposed that covering mine spoils with a peat blanket will actively 
prevent access of oxygen, as waterlogged peatlands are a reducing environment where organic matter 
accumulates as aerobic microbial decomposition is inhibited. In this anaerobic environment, a consortium 
of microorganisms reduce the plant cellulose to methane and carbon dioxide. Initial experiments are 
reported here to establish in the range and distribution of methane in natural bogs and the conditions 
necessary for the cultivation of methanogenic bacteria. An array has been set up in Mer Bleue near 
Ottawa, where the methane was sampled from three depths over an area of 24 m 2 . The amount of methane 
obtained from one station varied from 14 umoles to 10 mmoles, with a total accumulation from all stations 
of over 130 mmoles. Laboratory incubations appear to show that the methane measured in the field occurs 
where it is produced. It is postulated that the pore spaces of the peat are plugged by insoluble methane 
which limits movement of oxygen, water, or methane itself to provide an impermeable cover. 

Additional Key Words: methane, laboratory incubation, anaerobic environment, acid leaching. 

WETLAND/RIPARIAN RECONSTRUCTION FOLLOWING SURFACE MINING: PART III. RECOMMENDATIONS ON CHANNEL 
MORPHOLOGY TO SUPPORT A WET MEADOW 

Behling, R. E. (Professor, Department of Geology and Geography, West Virginia University, 
Morgantown, WV). Recommendations were formulated for reconstruction work on a 14-ha wetland/stream 
system, 927 m above sea level, which is being exposed through surface mining along Pendleton Creek, Tucker 
County', WV. This wet meadow is one of the numerous wetlands in the open-fold region of the Appalachian 
Plateaus in West Virginia and Maryland. The Pendleton Creek drainage basin is approximately 2.4 km wide, 
and the valley floor is 190 m wide. Fine-grained alluvium mantles the colluvium and bedrock to a depth 
of 2-3 m and forms the substrate for the natural wetland in the valley. The natural course of Pendleton 
Creek was a meandering channel, 4-5 m wide and 1.0-1.5 m deep. The gradient of the Pendleton Creek valley 
floor is less than 1 degree where the wet meadow exists. Sinuosity of the channel as determined from 
aerial photographs was 1.6. Bankfull width and depth measured in several locations between meander bends 
averaged 4.7 m and 1.4 m, respectively. Channel walls were nearly vertical with a curve at the base of 
the channel, creating a slightly concave floor. The natural channel is grass-lined. It is suggested 
that a grass-lined, meandering channel (sinuousity about 1.5), with rip-rap at the concave (outer) bank at 
meander bends would offer the greatest promise for all possible flow regimes of Pendleton Creek. The 
land adjacent to the channel could be returned to a wetland (a wet meadow would replicate natural 
conditions) and out-of-bank flow would be encouraged. A trapezoid cross-section for the channel of side 
slope 1.5:1; bottom width 3.0 m; depth 1.0 m would reflect bankfull conditions similar to those existent 
under natural conditions. 

Additional Key Words: stream hydrology, reclamation, stream flow path, controlled flooding. 

400 



A STAGED WETLAND TREATMENT SYSTEM FOR MINE WATER WITH LOW pH AND HIGH METAL CONCENTRATIONS 

Demko, T. M. (1) and Pesavento, B. G. (2). ((1) Geologist, P&N Coal Co. Inc., 240 West Mahoning St., 
Punxsutawney, PA. 15767 and (2) Consultant, Environment Analytic Service, Biological Filter Designers, 
Fredonia, PA). The use of man-made wetland ecosystems to renovate mine discharges has mainly been 
limited to low to moderate concentrations (10-100 mg/L) of iron, manganese, and aluminum. In order to 
study such a system under more adverse conditions, a wetland ecosystem complex has been designed and 
constructed to treat a mine discharge in western Elk County, PA, which is characterized by high 
concentrations of iron (100-400 mg/L), manganese (150-550 mg/L), aluminum (50-200 mg/L), and sulfate 
(8,000-12,000 mg/L), and a low pH (2.5~3.2). The flow rate varies seasonally and reacts quickly to 
precipitation and snow melt events, but averages 12 gal/min. The wetland ecosystem complex was designed 
to remove the high metal loadings by three mechanisms or groups of mechanisms: (1) the staged, bio- 
catalyzed oxidation/ precipitation of the metal ions and complexes in solution; (2) co-precipitation 
(adsorption, chelation, ion-exchange) of metal ions and complexes by iron, manganese, and aluminum oxides 
and hydroxides; and (3) inorganic geochemical precipitation. The wetland complex consists of six stages 
separated by flow and level control structures and aereation drop structures. The stages are, in down- 
stream order: an algae and moss pool, four separate typha-algae cells, and finally, a polishing pond. 
Preliminary results from the first year of treatment have shown that the acidity of the discharge has been 
reduced by two-thirds (2,500 mg/L to 850 mg/L) and the iron concentration has been reduced by three- 
fourths (400 mg/L to 100 mg/L). The transplanted and volunteer wetland vegetation have flourished in the 
lower two-thirds of the wetland during the first growing season. Metal precipitates have accumulated to 
depths of 1 .5 inches in the upper portion of the wetland. Flow has been reduced by one-fifth by 
evapotranspiration during the summer and early fall. Data from the second growing season are now being 
collected, and increased efficiency of treatment is expected. 

Additional Key Words: acid mine drainage, ecosystems. 



CONSTRUCTION OF A WETLAND DEMONSTRATION SITE FOR A METAL-MINE DRAINAGE 

Wildeman, T. R. (1), Laudon, L. S. (2), and Howard, E. A. (3). ((1) and (2) Department of Chemistry and 
Geochemistry, Colorado School of Mines, and (3) Department of Environmental Sciences and Engineering 
Ecology, Colorado School of Mines). At the Big Five Tunnel in Idaho Springs, CO, an EPA Superfund Site, a 
model wetlands ecosystem has been built. The project is sponsored by the EPA Region VIII Superfund 
Program and Camp, Dresser, and McKee under a contract to the Colorado School of Mines. Three 200 ft 2 
pilot plants have been constructed so the fate of all entering chemicals can be established. The 
objective of the project is to determine in one year whether wetlands can be used as a cost effective 
first step in the treatment of metal-mine drainages. So far the following conclusions have been made: 
1 ) The design was simple and the construction was straightforward (since all materials used were readily 
and locally available). 2) An area of 200 ft 2 was adequate to maintain a complete ecosystem. 
3) Selection of plants from local environments was possible, but transplanting was difficult. Large scale 
transplanting will require a significant investment in labor and equipment. 4) The mechanical portions of 
the system have operated well through a winter that was more severe than normal. 5) The chemical and 
bacteriological removal of pollutants has been occurring through the winter even though the vegetation was 
planted in September when it had already started to turn dormant. 

Additional Key Words: Superfund Program, acid mine drainage. 



RESPONSE OF THIOBACILLUS FERR00XIDANS TO ORGANIC COMPOUNDS: 
OF GROWTH 



LEAKAGE OF CELLULAR MATERIAL AND INHIBITION 



Bhatnagar, M. (1) and Singh, G. (2). ((1) and (2) Centre of Studies in Mining Environment, Indian School 
of Mines, Dhanbad - 826004, India). Chemolithotrophic bacteria, particularly Thiobacillus ferrooxidans , 
accelerate the oxidation of iron sulphide minerals. Materials toxic to these bacteria can be used to 
inhibit bacterial activity and control acid mine drainage (AMD). Iron oxidation by T. ferrooxidans , and 
growth on Fe 2+ iron were inhibited by a variety of organic compounds including sodium benzoate (SBZ), 
sodium lauryl sulphate (SLS), potassium sorbate (PSB), and some organic acids under laboratory conditions. 
Experiments were performed with cultures of AMD-bacteria to determine the threshold concentrations 
required for bacterial inhibition by candidate compounds. Results of this investigation indicated that 
SLS, SBZ, and PSB in the concentration range of 8 to 10 mg/L and organic acids in the concentration range 
of 10"3 M to 10~ 2 M, effectively inhibited bacterial growth with an associated decrease in iron oxidation 
of about - 70 percent. It is inferred that organic compounds either react with extracellular iron or affect 
the functions associated with the bacterial cytoplasmic membrane. 

Additional Key Words: acid mine drainage, sodium lauryl sulphate, sodium benzoate, potassium sorbate, 
lipopolysaccharide, peptidoglycan, cytoplasmic membrane. 



401 



IRON MONOSULFIDE AND PYRITE FORMATION IN SEDIMENTS OF LAKES THAT RECEIVE ACID MINE DRAINAGE 

Wicks, C. M. (1), Herman, J. S. (2), Mills, A. L. (3), and Schubert, J. P. (H). ((i) Graduate Student, 
(2) Assistant Professor, and (3) Associate Professor, Dept. of Environmental Sciences, Univ. of Virginia, 
Charlottesville, VA, and (H) Hydrogeologist, Argonne Natl. Lab, Argonne, IL) . Lake acidification from 
acid mine drainage (AMD) is an important environmental problem. Acid mine drainage can be ameliorated by 
the production of the minerals iron monosulfide (FeS) and pyrite (FeS 2 ) in impacted lake sediments. The 
authigenic formation of FeS 2 and FeS is hypothesized to occur when ferrous iron reacts with dissolved 
sulfide (H 2 S, HS"), which is produced by microbially mediated sulfate reduction. The result is that iron, 
sulfate, and acidity concentrations are decreased, while bicarbonate alkalinity is increased. If the 
reaction occurs in sediments that receive abundant sulfate ions and organic matter, then iron will be the 
limiting reactant. As the iron concentration increases, the dry weight percent of FeS or FeS 2 in the 
sediments should increase. The objective of this study was to determine if porewater iron concentration 
could be used as an indicator of FeS and FeS 2 formation in the sediments. Samples of lake water, 
porewater, and sediment were taken from 1H lakes that receive AMD throughout the Appalachian, Eastern 
Interior, and Western Interior Coal Basins. Concentration ranges observed in the porewater samples were 
0.1 to 1,500 mg/L Fe 2+ , 0.25 to H.2 mg/L Fe3 + , 2.6 to 1,850 mg/L SOj, 2 " , and pH of 6.05 to 8.0. Values of 
Eh calculated from the ferrous-ferric redox couple showed oxic conditions existing in the porewater. 
However, the sediments were dark in color and smelled of hydrogen sulfide, indicating anoxic conditions. 
The porewater composition, a reducing Eh, and field pH were used in the computer model PHREEQE to predict 
the saturation state of the porewater with respect to FeS 2 and FeS. Based on these results, in 9 of the 
14 lakes pyrite is predicted to be the stable mineral phase. The remaining 5 lakes are significantly 
undersaturated with respect to pyrite. Authigenic pyrite is not expected to be found in those sediments. 
For each of these lakes, the reaction is limited by low concentrations of either iron or sulfur. Molar 
Fe/S ratios in FeS 2 and FeS are 1:2 and 1:1, respectively. All undersaturated waters exhibit extreme 
departures from these ratios. 

Additional Key Words: porewater composition. 

HYDROLOGY AND GEOCHEMISTRY OF SURFACE COAL MINE LAKES 

Schubert, J. P. (Hydrogeologist, Argonne National Laboratory, Argonne, IL). Lakes are commonly created 
in surface-mined areas when mine pits or low depressional areas in spoil materials are not completely 
backfilled and eventually fill with water after mining operations have ceased. These lakes often contain 
good-quality water and have great potential post-mining benefits, such as swimming, boating, fishing, 
wildlife habitat, and even drinking water supplies. Some lakes, however, are very acidic (e.g., pH < H.0) 
and contain deleterious concentrations of sulfate, iron, manganese, and other trace metals. A study was 
initiated in an attempt to determine the hydrologic and geochemical characteristics of mine sites that may 
be controlling water chemistry in the manmade lakes. Twenty-one lakes in the Eastern and Central U.S. 
have been intensively sampled. Five lakes had pH values between 2.5 and 5.0, three lakes had pH values 
between 6.0 and 7.0, eight lakes were between 7.0 and 7.5, and five lakes had pH values greater than 7.5. 
Hydrologic variables such as infiltration rates, hydraulic conductivity, and texture of spoil materials 
have been measured. Geochemical analyses of soil, spoil, and lakebed samples included: pH, electrical 
conductance, cation exchange capacity, exchangeable cations, total sulfur, and neutralization potential. 
Multivariate statistical procedures have been utilized to determine which watershed characteristics, 
hydrologic variables, geochemical parameters, or lake morphometric variables are the most influential on 
the resulting water chemistry of mine lakes. 

Additional Key Words: acidity, trace metals, watershed characteristics. 

WATER RESOURCE DEVELOPMENT ENGINEERING AND ACID MINE DRAINAGE IN THE UPPER OHIO RIVER BASIN 

Koryak, M. (Limnologist , U.S. Army Corps of Engineers, Pittsburgh District). Acid mine drainage from 
bituminous coal mines has been the greatest single water pollution problem in the upper Ohio River Basin. 
Thousands of miles of streams within the western Pennsylvania, northern West Virginia, western Maryland, 
and southeastern Ohio portions of the basin have been degraded by an acid mine drainage (AMD) load that, 
until recent decades, was equivalent to more than a million tons/year of sulfuric acid. Severe AMD 
pollution caused damage by corroding pipes, pumps, boats, gates, and navigation aids. The acid and 
associated mineralization, and the frequent gross heavy metal pollution degraded the aesthetic and 
recreational value of local waters. The AMD suppressed and often totally eliminated aquatic life in local 
impoundments and along substantial reaches of major rivers, and it caused numerous and serious domestic 
and industrial water supply problems. Because of the extent and magnitude of the problem, AMD 
considerations have had a significant influence on many aspects of water resource development engineering 
in the upper Ohio River Drainage Basin. In some areas, AMD necessitated the use of special construction 
techniques and corrosion-resistant materials, and it increased and complicated maintenance problems. It 
was a major influence on the planning, design, and operation of large civil works engineering projects 
constructed prior to passage of the 1972 Clean Water Act. Reservoir operation schedules were developed as 
integral parts of Corps of Engineers reservoir projects to moderate low-flow AMD degradation extremes. 
As dramatic progress has been made recently in AMD abatement, these projects and operations continue to 
provide, for the most part, very substantial AMD mitigation benefits. 

Additional Key Words: heavy metal pollution, water resource development, upper Ohio River Basin. 

402 



AERATION EFFICIENCY OF THE IN-LINE AERATION AND NEUTRALIZATION SYSTEM 

Ackman, T. E. (1) and Hustwit, C. C. (2). ((1) Mining Engineer, U.S. Department of the Interior, Bureau 
of Mines, Pittsburgh Research Center and (2) Project Engineer, Boeing Services International). Acid mine 
drainage treatment typically involves the removal of soluble metals by neutralization, oxidation, and 
precipitation. The chemical reactions are frequently limited by the availability of dissolved oxygen. 
For this reason a study was conducted in the laboratory to establish the aeration efficiency of the In- 
Line Aeration and Neutralization System (ILS). This patented treatment system, developed by the Bureau of 
Mines, treats mine discharge in a pipeline and involves no moving parts. It consists of two components: a 
jet pump and a static mixer. Testing procedures were performed in accordance with the American Society of 
Civil Engineers (A Standard for the Measurement of Oxygen Transfer in Clean Water). The standard requires 
deoxygenation of a tank of clean water before aeration. The dissolved oxygen concentration in the water 
is measured over time as it is continuously pumped through the ILS until saturation is achieved. 
Corrections are made for ambient water temperature and barometric pressure. A total of twenty-eight tests 
were conducted using two different jet pumps and seven types of static mixers. Each combination of jet 
pump and static mixer was tested at two operating pressures: 20 psig and 50 psig. The results indicated 
that there were significant differences in aeration efficiency based principally on the water pressure. 
All systems transferred more oxygen when operated at 50 psig. Differences between the two jet pumps and 
four of the mixers do not appear to be significant. The Standard Oxygen Transfer Rates (SOTR) for the 
best jet pump/static mixer combinations are comparable to those cited by manufacturers of mechanical type 
aerators. 

Additional Key Words: oxidation, aeration, sulfite, water treatment, acid mine drainage. 

TREATMENT OF ACIDIC COAL REFUSE BY ADDITION OF CRUSHED-STONE POND SCREENINGS 

Stokowski, S. J., Jr. (1) and Gilbert, R. R. (2). ((1) Consulting Geologist, Ashland, MA and (2) Civil 
Engineer, Adger, AL). Acidic refuse from Warrior Basin (AL) coal was treated with carbonate-rock, 
crushed-stone pond screenings, resulting in an improvement in the pH from 2.9 to 6.6. Fe(0H)o (yellowboy) 
formed on the treated waste and in ditches. Consequently, the pH of the runoff water improved from a 
range of 2.2-2.6 to a range of 4.4-6.2. Aqueous Ca and Mg concentrations and specific conductivity also 
increased. The water color changed from red to blue-green, possibly due to the iron oxidation state. 
The improvements have persisted for over 2 years from a single application of 54 tons/acre. The treatment 
process consisted of: 1) grading and ditching the refuse, 2) dumping the pond screenings and spreading 
with a bulldozer, 3) discing the material into the upper 6 inches of refuse, and 4) monitoring pH and 
other chemical variables in the refuse and surface-water runoff. After a few months of stable, near 
neutral pH conditions, grass was planted and is still vigorous. The refuse is waste material from the 
processing of coal, and is 6 inches and finer in size. It consists of shale that often contains 
marcasite, shale interbedded with coal and nodules or iron-pyrites with attached coal. The surface of the 
pile readily weathers into an impermeable, fine-grained layer. The pond screenings used are a byproduct 
of limestone/dolomite aggregate (Knox Group, AL) washed for concrete and other uses. They are especially 
finely graded (98? passing #50 sieve, 87? passing #200 sieve, 80? passing #325 sieve) because coarser 
fractions are removed by hydrocyclone for use in other products. The finest fractions are discharged to a 
sedimentation pond and recovered by dragline. Stockpiling and free drainage reduces the moisture content 
to acceptable levels. The fine gradation and chemical composition (79? CaCOo, 14? MgCOo) make the product 
an efficient neutralizer of coal mine acids. 

Additional Key Words: reclamation, limestone, dolomite. 

X-RAY DIFFRACTION EVALUATION OF COAL OVERBURDEN NEUTRALIZATION POTENTIAL 

Despard, T. L. (Geologist, Office of Surface Mining Reclamation and Enforcement, Knoxville, TN). The 
Federal Regulations established to implement the Surface Mining Control and Reclamation Act of 1977 
(SMCRA) require the identification of acid/toxic overburden strata and coal and the determination of the 
potential for such materials to adversely impact ground- and surface-water resources. The net acid-base 
account method is widely used to determine the acid/toxic potential of coal-bearing strata. This method 
is unreliable when applied to noncalcareous rocks because in these, the analytically determined 
neutralization potential is primarily due to silicate minerals, which react slowly with acid generated 
from pyrite oxidation. Leach testing has been used as an alternative to the net acid-base account method 
as a predictor of acid-mine drainage production; however, these tests are expensive and time-consuming. 
Determining the carbonate mineralogy of overburden strata, in conjunction with the net acid-base account, 
can greatly assist in predicting the effectiveness of the measured neutralization potential. X-ray 
diffraction (XRD) analysis can determine carbonate mineralogy both on a qualitative and semiquantitative 
basis, and the analyses can be accomplished rapidly. For the carbonates, calcite, dolomite, magnesite, 
and siderite, the required scan range is only from approximately 20° 29 to 33° 29, and can be accomplished 
in minutes. Computer routines are available that can compute the percentage of calcite, dolomite, 
magnesite, and siderite in rock samples and can be expanded to include other mineral species if desired. 
These routines are derived from the generalized basic equation of X-ray quantitative analysis: 

Ij = IjCijXj i = 1 ,2 m, where c^j = K^j/pjuT 

Additional Key Words: overburden mineralogy, AMD prediction, acid-base accounting, carbonates. 

403 



STREAM SEALING OVER AN ACTIVE LONGWALL PANEL 

Ackman, T. E. (1) and Huatwit, C. C. (2). ((1) Mining Engineer, U.S. Department of the Interior, Bureau 
of Mines, Pittsburgh Research Center and (2) Project Engineer, Boeing Services International). Stream 
infiltration losses associated with longwall mining activities can initiate environmental and mining 
hazards. In order to identify potential stream loaa zones that might occur aa a reault of active longwall 
mining, the Bureau of Mines in cooperation with Southern Ohio Coal Company /Martinka Division instituted a 
study of surface water flow along a 1,400-ft section of Guyaes Run, located near Fairmont, WV. Seven 
stream gaging and forty electromagnetic conductivity atationa were eatablished in July 1987, about 3 
months before passage of a longwall panel beneath the stream section. Analysia of aurvey data collected 
within the stream channel permitted identification of a 300-ft zone of higheat probability for stream 
infiltration losses. Conductivity surveys of the fracture syatem beneath the adjacent flood plain alao 
confirmed the location of a potential infiltration loaa zone in the atream channel. During the passage of 
the longwall panel beneath the predicted 300-ft loss zone, a complete stream loaa event occurred within 
this section. Total water infiltration into the mine was estimated to be between 500 and 700 gal/min. 
Immediate steps to restore the streamflow involved sealing the channel with high-pressure injections of 
polyurethane grout. Approximately 11,800 lb of grout were injected through 153 grout roda located within 
the stream channel loas zone and along the contiguous channel banks. The rod emplacement depths varied 
between 3 and 6 ft. Within a few hours after completion of grouting, surface water flow was reatored to 
the stream. This streamflow restoration virtually eliminated water infiltration into the mine. Post- 
grouting conductivity and stream gaging meaaurementa have ahown a flow pattern conaiatent with premining 
activities, and water infiltration into the mine has continued to be minimal. 

Additional Key Worda: electromagnetic conductivity, atream infiltration, atream losses, polyurethane 
grout. 



MICROMAP -- A DATABASE RETRIEVAL AND DISPLAY MANAGER FOR MICROCOMPUTERS 

Rymer, T. E. II (1), Renton, J. J. (2), and Stiller, A. H. (3). ((1) Research Associate, Departments of 
Chemical Engineering and Geology, West Virginia University, (2) Professor, Geology Department, West 
Virginia University, and Geochemist and Head of Analytical Section, West Virginia Geological and Economic 
Survey, and (3) Associate Professor, Chemical Engineering Department, West Virginia University). A new 
software package has been developed called "MICR0MAP" to expand the capabilities of the Coal Reclamation 
Information System (CRIS) computer software developed in an earlier project funded by the West Virginia 
Dept. of Energy through the West Virginia Geological Survey. The CRIS software allowed the user to 
extract information from a database containing general mine information, water quality data, overburden 
analyses, and reclamation-revegetation data for surface mines in 12 northern West Virginia countiea uaing 
a variety of geographical related retrieval modea (county-wide, 7.5 minute quadrangle, radiua of a desired 
coordinate, and watershed). Unlike the original mainframe software, MICR0MAP is able to run on any IBM/PC 
compatible microcomputer. The computer displays a map of any county(ies) requested by the user. The 
locations of surface mines that are in file, overburden drill core locations, and water sampling locations 
are pinpointed on the map. Data are extracted by moving a small graphics cursor to any location on the 
map and designating a square mile search area. The computer will then retrieve the general mine 
information, water quality data, and overburden analysis data for the designated area and move it to a 
display manager where it can be viewed on the screen and, if desired, channelled to a hard copy printer. 
In addition, a 7.5 minute quadrangle search mode is available, where the user need only specify a desired 
quadrangle. The original CRIS data sets have been subsetted into smaller, county data aeta. The 
original master data set still resides on the mainframe where it can be updated and revised. The 
mainframe computer system generates the smaller county-wide data sets and moves these data sets to floppy 
disks. There are neither statistical nor graphical options available with MICROMAP as it is primarily a 
data search and retrieval display manager. Its chief advantage is its mobility and ready, easy access 
requiring only a microcomputer. The map graphics are highly accurate. It is possible to display Preston 
County, WV and Fayette County, KY on the same screen and still retain very good resolution and search 
capabilities. The mapping algorithm used is an efficient modification of the gnomonic method of map 
projections. MICROMAP has the capabilities of displaying all 55 West Virginia counties and corresponding 
datasets with minimized distortion on a medium resolution IBM color graphics screen. 

Additional Key Words: mine sites, water quality, water sampling, CRIS, West Virginia. 



404 



INSTALLATION AND STABILITY OF INVERTED PYRAMID-SHAPED PLUGS 
FOR CLOSING ABANDONED MINE SHAFTS GALENA, KS DEMONSTRATION PROJECT 1 



John S. Volosin 2 



Abstract.-- The Bureau of Mines designed and installed 11 
inverted- pyramidal- shaped plugs in a mine closure 
demonstration project completed in Galena, KS in December 1983. 
The demonstration project resulted from a study done by the 
Geological Surveys of Missouri, Kansas, and Oklahoma for the 
U.S. Bureau of Mines in January 1983. This study identified 
over 1400 open mine shafts and nearly 500 subsidence collapse 
features that remained from the original 14,000 shafts sunk in 
the Tri-State Zinc-Lead Mining District. In Galena, KS, alone 
over 377 open mine shafts were readily accessible and 150 
abandoned mine shafts were within the city limits. Of these, 
the Bureau of Mines selected 14 abandoned sites for the mine 
shaft closure demonstration project. During the demonstration 
project, 11 mine shafts were closed with the inverted pyramid 
shaped reinforced- concrete plugs, 2 were closed with 
reinforced concrete caps after backfilling, and one shaft was 
closed by backfilling only. The stability of the closure 
devices has been monitored and evaluated over a 3-year period. 
The results indicate that the Bureau of Mines closure devices 
are stable and have eliminated hazards associated with open mine 
shafts in a populated area. 



Introduction 

The Bureau of Mines operated a shaft closure 
demonstration project in Galena, KS, in which 14 
abandoned mine shafts were closed. The primary 
purpose of the project, which was part of the 
Bureau's program for conserving land resources, was 
to provide alternative methods for closing some of 
the hundreds of open shafts in the Tri-State area. 
The project developed from a study to evaluate the 
hazards of the abandoned zinc-lead district. 



The Galena field is in the Tri-State zinc-lead 
belt district of Kansas- Missouri- Oklahoma, which 
was one of the largest zinc-lead mining districts 
in the country. The district produced over 11 
million tons of zinc and 2.8 million tons of lead 
during its 122 years of operation. The total value 
of the lead and zinc produced in the district from 
1850 through 1970, in terms of recoverable metal, 
was $2,073,200,000. In terms of today's dollars, 
the value would be in the neighborhood of 20 
billion dollars (Stewart 1986, Martin 1946) 



ipaper presented at the 1988 Mine Drainage and 
Surface Mine Reclamation Conference sponsored by 
the American Society for Surface Mining and 
Reclamation and the U.S. Department of the Interior 
(Bureau of Mines and Office of Surface Mining 
Reclamation and Enforcement), April 17-22, 1988, 
Pittsburgh, P. A. 

Metallurgist, Rolla Research Center, Bureau 
of Mines, Rolla, M0. 



405 



Mining began in the Tri-State district in 1848 
with the discovery of ore deposits in Joplin, Mo. 
The deposits were originally mined for their lead 
value, but the development of the railroad in the 
area and the coincidental development of the new 
milling and smelting techniques for zinc resulted 
in the area becoming a valuable source of zinc. 
Zinc production in the district began with the 
first shipment of concentrates to LaSalle, IL in 
1872. Later shipments went to the smelter in Weir 
City, KS constructed in 1873. By 1875, the Joplin 
field became the leading zinc producer in the USA. 
Additional lead-zinc deposits discovered west of 
Joplin increased mining activity. The Galena, KS, 
field was discovered in 1877. In 1891, lead mining 
began in Indian Territory (northeast Oklahoma), 
near Peoria, and ore discoveries followed near 
Lincolnville, Miami and Picher. The Commerce, OK, 
field was discovered in 1905. Large scale mining 
started in the Miami-Picher, OK, field in 1916. 
Mining continued in the Missouri portion of the 
district until 1957 and in the Kansas-Oklahoma 
portion until 1970. (Dressel et.al. 1986, Fejes 
et.al 1985, U.S. EPA 1975) 

Early mining leases were generally small with 
many leases being only 100 or 200 ft 2 (Clerc 1907, 
Hay 1893, Hayworth 1901, Norris 1968, Plyn 1904). 
The ore was mined by small crews of men using hand 
tools and simple hoisting devices. Exploration was 
done by sinking a shaft, generally 50 to 100 ft 
deep, until ore was found. Exploration continued 
by drift mining outward from the shaft (Crane 
1901). If ore was not encountered, the miner moved 
to new ground and sank another shaft. The mines 
were generally developed with little regard to any 
long-range overall mine plan. If drifts reached 
300-ft in length or if ventilation became 
difficult, additional shafts were sunk. In most 
cases, the underground mine workings were not 
mapped. 

In Galena, KS, the mining depths varied from 
ground surface to 300 feet. The ore was generally 
confined to thin bedded strata, and ore bodies were 
usually less than 30 feet thick. But, in some 
cases, ore zones 80 to 100 feet in thickness were 
developed. The ore was worked from the upper 
portion of the ore down to the depth at which water 
became too much of a problem to continue mining. 
When pumping facilities were installed, some of 
these mines were reopened and mining of the thick 
ore deposits resulted in rooms as much as 100 feet 
in height. In 1893, Henrich reported that diamond 
drills were being used for prospecting deeper than 
100-ft. By about 1900, the churn drill replaced 
shaft sinking as the principal exploration tool 
(Gibson 1972, Plyn 1904). 

The use of shafts as a means of exploration 
and the small lease and subleasing of mining plots 
resulted in a high density of mining shafts in the 
area. In preparing a series of reports for the 
Bureau, the State Geological Surveys of Kansas, 
Missouri, and Oklahoma located over 1,400 abandoned 
open mine shafts remaining of the original 14,000+ 
prospect and mining shafts sunk in the Tri-State 
district (Luza 1983, McCauley et.al. 1983, 
McFarland and Brown 1983). Although 90% of the 
original number of shafts have been closed, the 
shafts that remain open are constant safety and 
environmental hazards and limit the use of the 
land. In the Galena, KS, field alone, 377 open 
shafts were located within or adjacent to the city 
limits of Galena. All but 11 of these shafts 



showed surface enlargement because of cribing 
removal or failure. 

Several methods have been used for closing 
shafts (Genie Eng. LTD 1983, NCB 1982). In the 
Tri-State shafts have been closed by backfilling 
and capped with timber caps, steel plates, concrete 
slabs, and railroad rail- gratings. However, in 
some cases the closure device failed and the shaft 
reopened to again become a safety hazard. 

Backfilling was a common method for filling 
shallow shafts and it is still quite a successful 
method if done properly with graded material free 
of degradable trash and in a manner that avoids 
temporary bridging. Timber caps, rails, and steel 
plates have been used with varying degrees of 
success but eventually decay or rust, resulting in 
an unsafe closure. Although concrete caps have 
been successful in some instances, there are 
examples of failed concrete caps in the Galena, KS, 
area where the caps were improperly reinforced or 
where washout caused the cap to tip on end. In the 
Picher, OK, field, at least one company 
successfully used concrete cubes to close shafts 
when abandoning the field. The cubes, which were 
6-1/2 ft on a side, were constructed on the surface 
next to the shaft and then rolled into the opening 
and wedged into place by undercutting and blasting. 

Three methods were used for closing abandoned 
mine shafts during the Bureau's demonstration 
project described in this report. The newest 
method, the installation of the inverted 
pyramid-shaped plug designed by Bureau personnel, 
is discussed in detail. The results of 3- years of 
monitoring are also included. Additional details 
are available (Dressel and Volosin 1985). 



SELECTION OF SHAFT SITES 

Galena, KS, was picked as a site for 
demonstrating methods for closing abandoned shafts 
because of the large number of open shafts that 
were readily accessible. Before a location was 
selected for the demonstration, the Bureau 
contacted the Galena city government to ascertain 
which areas within the city limits were in the most 
need of shaft plugging. From the locations the 
city officials listed, the Bureau selected a site 
in NE1/4SW1/4 sec. 14, T. 34 S., R. 25 E., at the 
west end of Front, First, and Second Streets. 
Virtually all of the open mine shafts in Galena are 
on privately owned land. A search of the county 
land records was made to determine the ownership of 
the lands and the owners were contacted to obtain 
grants of easement that would permit the Bureau to 
carry on the demonstration project. Figure 1 shows 
location of the site and the location of the shafts 
closed during the demonstration. The breaks in the 
shaft numbering sequence resulted from changes in 
the original closure plan, brought about in one 
case because considerable construction debris had 
been dumped into the opening, and other cases 
because the locations were out of the area covered 
under the grants of easement. 



406 



Two of the open shafts selected for plugging 
are shown in figures 2 and 3. A contract was let 
in 1982 to obtain the shaft dimensions, assess the 
conditions of shaft side walls, obtain the 
elevation of an estimated contact between 
overburden and solid rock, and the location of 
underground workings in the shaft vicinity. These 
measurement were used in preparing the competitive 
bid specifications for the plug installations 
reported herein. Another contract was awarded in 
August 1983, to perform the actual closure 
demonstration work in which 11 shafts were plugged, 
2 were cappped, and 1 was backfilled. 





^7 '2 






i 3 ■ ii 




^14 


■ ■ 




■ 


»■ .10 




'■ 






2 a 


'■_6 




■* 


■is a ' 












Figure 1.— Project site location. 








Figure 2. —Typical shaft showing rotted timber and washout along timbers, 




Figure 3. --Typical open shaft showing circular outline in slumping residuum. 



407 



DESIGN OF PLUG 

An inverted pyramidal design was selected 
because it fulfilled the following criteria: (1) 
simplicity of construction, (2) ease of 
installation, (3) personnel safety during 
installation, and (4) permanency of the installed 
closure device. The shafts in this demonstration 
were roughly square and ranged in size from 4 to 8 
ft. The pyramidal plug design was chosen because 
it was easily adapted to a variety of irregular 
mine shaft openings with a minimum amount of site 
preparation. When installed in a shaft, the center 
of gravity of the inverted pyramid plug could be 
placed so that the plug would have a tendency to 
adjust and wedge tighter into the shaft. A 
lightweight, prefabricated disposable form was 
designed that required a minimum amount of time to 
prepare and set in place without having to use 
heavy construction equipment. The form, complete 
with concrete reinforcement rods, was constructed 
in a welding shop away from the demonstration site 
and installed with a 17-ton crane. 




Figure 5. — Eight-ft form showing reinforcing grid. 



The forms were designed to be totally 
self-supporting. Once they were set in place they 
were not anchored to any other structure within the 
mine shaft. They were constructed in three 
standard sizes of 8, 10, and 12 ft. The size for a 
given shaft was selected so that the top of the 
plug was approximately 4 ft larger on a side than 
the size of the shaft opening. The reference 
monument, a 4-inch pipe, long enough to extend 
above the surface level of the ground after 
backfilling, was attached to the center of the 
plug. A sketch of an installed plug is shown in 
figure 4. 




S 

E 

C 

T 

I 


N 



Figure 4. --Installed plug. 

Eleven forms were required for the 
demonstration: 3 were 8-ft by 8-ft, 6 were 10-ft by 
10-ft, and 2 were 12-ft by 12-ft. The 8-ft by 8-ft 
and the 10-ft by 10-ft forms were constructed of 
3/16-inch hot-rolled low-carbon steelplate welded 
at the seams. The 12-ft by 12-ft forms were 
constructed of 1/4-inch hot rolled low-carbon 
steel . The external edges of the seams were 
reinforced by the addition of a fabricated angle, 
approximately 3-inches on an edge, welded to the 
seam. This reinforced edge proved very beneficial 
since much of the weight of the plug rested on the 
corners before seating. A plate was welded in each 
corner with an eyelet for attaching cables for easy 
handling and positioning the forms (fig. 5). 



A horizontal reinforcing rod grid was placed 
1 -ft from the top in the 8-ft pyramid forms, 
1.25-ft from the top in the 10-ft forms, and 1.5-ft 
from the top in the 12-ft forms. A 12-inch grid 
spacing was used as shown in figure 5. Grade 60, 
No. 7 reinforcing bars were used in each instance. 

In the 10-ft form, two S4 I-beams with 
0.326-inch web thickness approximately 6 ft long, 
were welded to a 1/4- inch footplate which was 
welded to the sides of the form. The beams were 
arranged parallel to each other and spaced 
approximately equidistant from the parallel side 
walls and from each other (fig 6). 




Figure 6.— Ten-ft form showing I-beam position 
arid reinforcing grid. 



To brace the side walls of the 12-ft form, two 
S4 I-beams, approximately 8 ft lopq with a 
0.326-inch web thickness, were welded to the form 
at right angles to each other. The ends of these 
beams were welded to a 1/2-inch plate at least 1 
ft', which in turn was welded to a second 1/2-inch 
plate at least 1 ft 2 , previously welded to the 
inside of the form approximately 2-ft from the top 
(fig. 7). 



408 



Vertical re 
from each of the 
form (fig. 7) . 
top of the form 
near the bottom 
extended to with 
A spacer was wel 
of each form to 
side walls. 



inforcing bars were placed 6 inches 
sloping sides of the pyramidal 
They were spaced 1-ft apart at the 
and tapered down to a few inches 

The top end of the rebars 
in 6 inches of the top of the form, 
ded approximately 1-ft from the top 
hold the rods 6 inches from the 




Figure 7.--Twelve-ft form'showing I-beams and 
position of reinforcing bars. 



Material requirements per plug are tabulated 
in table 1. 

Table 1. --Material requirements per plug. 

Approximate shaft size 
4-ft 6-ft 8-ft 



Pyramid form size ...ft,. 8x8x4 10x10x5 12x12x6 

Metal preform ft 2 . 1 90.51 1141.2 2 203.6 

Weight of metal form.. lb. 693 1,083 2,079 

Estimated rebars per shaft: 

Linear ft 327 495 698 

Weight lb.. 660 1,020 1,420 

I-beam: 

Length ft.. 12 16 

Weight lb.. 89 118 

Edge angle: 

Length ft.. 27.7 34.6 41.6 

Weight lb.. 260 325 391 

Total weight of steel. lb. 1,613 2,517 4,008 

Concrete yd 3 . 3.2 6.2 10.7 



^3/16-inch cold-rolled, low-carbon steel plate. 
2 l/4- inch cold-rolled, low-carbon steel plate. 

INSTALLATION OF PYRAMIDAL- SHAPED PLUGS 

A minimum of site preparation was required to 
prepare the shafts for plugging. The contractor 
used an Extenda-Hoe backhoe to remove sufficient 
material from around the surface of each open shaft 
to provide a roughly level contact of the 
disposable form with the interface between the 
surface residuum and the bedrock. The edges of the 
shaft at the interface were trimmed to allow the 
center of gravity of the form to be set below the 
interface elevation and to have a suitable bearing 
surface for setting the forms. Trimming was done 
with a jackhammer attached to the backhoe. 



The pre-fabricated forms were delivered to the 
site on a flatbed trailer and were unloaded from 
the truck and placed directly in the hole using a 
17-ton crane (fig. 8). In several cases, 
additional sidewall trimming using the backhoe 
and/or jackhammer was required to obtain a level 
postion for the forms. 




Figure 8. --Lowering 10-ft form into opening. 

Class A concrete was delivered to the site 
from a local batch plant. Just prior to pouring, a 
reference monument was attached to the center of 
each plug (fig. 9). The 8-ft forms each required 
3.2 yd 3 of concrete, the 10-ft forms 6.2 yd 3 , and 
the 12-ft forms 10.7 yd 3 . 

After the first three plugs were placed, it 
was noted that there were gaps between the form and 
the sidewalls along the edge of the pyramid shaped 
form. In these areas, reinforcing bars were 
positioned over the edge of the form and extended 
to the side of the prepared opening. This was 
covered with a 2-ft width of expanded metal, and 4 
inches of concrete was poured on this expanded 
metal . 




zffi 



^{V 



-if. 



m 



%$*mr. 



.*-- 



;4v 



7? 



j&^jCJ 



i-Af 










Figure 9.--Ten-ft form showing sjyder-leg reinforcing 
over edge of form to support the expanded metal . 



409 



Before pouring co 
forms, steel reinforci 
to the installed horiz 
extended in spider-leg 
the edge of the form ( 
expanded metal was pos 
reinforcing rods, and 
when the form was fill 
1 to 1-1/2 yd 3 of cone 
shaft because of these 



ncrete in the last eight 
ng rod was bent and fastened 
ontal reinforcing grid and 

fashion several feet over 
fig. 9). The 2-ft wide 
itioned over these 
concrete was poured over this 
ed (fig. 10). An additional 
rete was required for each 

modifications. 



& 






i 








Figure 10.--Ten-ft form filled with concrete. 

In the installation of one 12-ft and one 10-ft 
plug, the center reinforcing bars were left 
unwelded to the sides of the form to allow the 
form to bulge along the edges to better fill the 
gaps. This appeared to be an effective measure. 

After a concrete curing period of at least 7 
days, the excavated areas were backfilled with 
waste rock available near the shafts. The backfill 
at each site was mounded so that the center of the 
fill was approximately 2 ft above the surrounding 
ground surface. The 4-inch pipe extending from the 
center of each plug was trimmed so that it extended 
6 inches above the fill; it was filled with 
concrete and was designed to remain as a marker for 
evaluation purposes. 

INSTALLATION OF SLAB CAPS 

In some instances, solid rock exposed at the 
surface was competent enough so that there was 
virtually no cratering or shaft enlargement at the 
surface. In these cases, it was found expedient to 
trim away loose surface rock and install reinforced 
concrete slabs. Two mine openings, approximately 
4.5-ft by 4.5-ft, were closed by this method during 
the completion of the project. 

The slabs were designed to extend 
approximately 5-ft over each edge of the open mine 
shafts. A reinforcing grid of No. 7 rebars was 
installed. The bars were spaced 1 -ft apart at the 
edges, 1/2 ft apart in the area over the shaft 
opening, and located 1/2 ft from the bottom of the 
pad. A monument pipe was attached to the center of 
each grid, and an 18 inch thick pad of class A 



concrete was poured. The material requirements for 
each poured slab are shown in table 2. 

Table 2. --Material requirements per slab. 



Size of shaft ft.. 4.5x4.5 

Dimension of slab ft.. 15x15x1.5 

Number of No. 7 rebars 46 

Total length of rebars ft.. 690 

Total weight of rebars lb.. 1,420 

Thickness of concrete ft.. 1.5 

Volume of concrete yd 3 .. 12.5 



These slabs are of sufficient size 
and sufficiently reinforced to remain 
indefinitely without breaking or flipping 
over in the event of washout under part of 
the cap. They were also designed to 
withstand loads from automobile or truck 
traffic which may occur in the area 
following closure of the shafts. 

INSTALLATION OF BACKFILL 

In preparing shaft Mo. 14 for plug 
installation, the hole was found to be 
larger than original surface measurement 
had indicated, and it was difficult to 
obtain a stable bedrock surface. It 
became apparent that th3 alternatives were 
to either install a 16 ft plug, a 22 ft 
cap, or completely backfill the hole. 
Backfilling was chosen. This method, 
currently the locally accepted way to 
close a shaft, proved useful to this 
project in that it provided a reference 
for evaluating the closing of shafts by 
either plugging or capping abandoned mine 
shafts. Approximately 350 yd 3 of backfill 
were required to close the opening. A 
4-inch pipe was placed at the shaft center 
as a marker for this closure. 

DISCUSSION OF CLOSING DEMONSTRATION 

A minimum of installation problems 
were involved and only a few changes were 
necessary in completing the planned 
demonstration program. The most serious 
problem was that no suitable underground 
maps were available for the demonstration 
site. 

Working around abandoned mines 
shafts, particularly in areas where the 
extent of the abandoned underground mine 
working and the stability of the surface 
material around the shafts is not known, 
is potentially hazardous. Most of the 
abandoned mines are filled or partially 
filled with water. Water, encountered in 
all shafts during the mine shaft 
inspection performed by the contractor, 
made it difficult to determine if either 
the actual bottom of the shaft or a 
temporary bridge had been reached. Some 
shafts were filled with debris, such as 
old autos, air conditioners, and/or 
refrigerators. There was no way to get 
around or through debris at the bottom of 
the shafts. No debris was removed during 
the closure demonstration. 



410 



It also became obvious that the shaft 
dimension measurements made from the 
surface in these old hand-dug shafts 
before they were prepared for plugging 
were not entirely reliable. In several 
instances, measurements, made after the 
holes were prepared, required changes in 
plug size from original specifications. 
For example, the plug put in open shaft 
No. 19 was the plug originally constructed 
for shaft No. 14. 

Sufficient waste rock was available 
on the nearby surface to backfill the 
openings over the installed plugs; most of 
the material trimmed from the shaft 
openings was allowed to fall into the open 
shafts. In one instance, temporary 
bridging of a shaft occurred during the 
trimming operation. This was to be 
expected because the material was 
ungraded, but it pointed out the necessity 
for using graded material when closing 
small shafts by backfilling. 

The project was completed using a 
minimum amount equipment at the project 
site. Required equipment was a backhoe, a 
flatbed truck, a crane, and a concrete 
delivery truck. Care was taken to 
strategically locate the equipment around 
the shaft collar to avoid parking the 
vehicles above the underground mine 
workings and no more than two vehicles 
were at the shaft at anytime. The 
contractor and suppliers were informed of 
the possible hazards and proceeded with 
due caution and regard for personal 
safety. 

The reinforcing rods were welded into 
position at crosspoints, and the ends were 
welded to the sides of the forms. This 



made the forms very rigid. The welding of 
the reinforcing rod ends at the center of 
the sides was omitted in several of the 
forms installed near the end of the 
demonstration. This enabled the form to 
bow out when filled with concrete to more 
nearly take the shape of the opening. 
However, when this happens, there are no 
reinforcing bars in the bowed part of the 
plug. A modified design that allowed the 
center reinforcing bars to extend through 
the plug walls might eliminate this 
problem. 

Four plug forms were set in their 
respective shafts prior to a 7-inch 
rainfall. As a result, the forms were 
filled with water and had to be pumped out 
before the concrete was poured. Small 
drain holes were left in the last three 
plugs installed; however, the plugs were 
filled with concrete before the next rain. 

EVALUATION OF CLOSURE DEVICES 

After the 14 shafts were closed, the 
area around each closure device was 
backfilled to the rough surface elevation 
of the immediate area. Once this rough 
grade was established for the 
demonstration site, a 2-ft mound of fill 
was formed directly over each shaft and 
the tops of the reference markers for each 
shaft were trimmed to approximately 
6-inches above the fill material. A local 
registered land surveyor was hired to 
establish a horizontal and vertical survey 
control net within the mine closure 
demonstration area. The initial elevation 
of the reference markers and the results 
of resurveys each year for a 3 year period 
are shown in Table 3. 



Table 3. --Elevation of reference point on each closure device. 



CLOSURE DEVICE 
In Shaft Number 



Survey results showing reference point elevation, ft 
12/23/83 10/31/84 9/27/85 11/14/86 



8 FT PYRAMID PLUGS 

2 900.59 900.59 900.59 

4 898.46 898.44 898.44 

9 901.98 901.97 901.97 

10 FT PYRAMID PLUGS 

5 890.74 890.74 890.73 

7 897.49 897.47 897.47 

8 887.36 887.35 887.35 

11 901.56 901.54 901.54 

12 897.01 897.01 896.99 

19 890.49 890.47 890.46 

12 FT PYRAMID PLUGS 

6 892.34 892.33 892.33 

10 902.76 902.70 902.70 

15 FT CONCRETE CAP 

1 901.90 901.90 901.90 

13 895.12 895.09 895.08 

BACK-FILLED ONLY 

14 898.21 897.95 ^caved 

kaved 23 ft, estimated 125 yd;* to reestablish to grade elevation. 

2 Caved 30 ft, estimated 150 yd 3 to reestablish to grade elevation. 



900.59 
898.44 
901.97 



890.72 
897.47 
887.35 
901.54 
896.99 
890.46 



892.33 
902.70 



901.90 
895.08 



^caved 



411 



All the closure 
period of 12 months 
noticed around shaft 
demonstration area 1 
the demonstration wo 
noted around any of 
of shaft No. 14 a mo 
reference monuments 
shaft No. 14 had cav 
about 125 yd 3 of fil 
grade. The fill in 
subside. The survey 
that the crater woul 
to reestabl ish grade 
any of the other clo 
after 3 years of mon 



devices remained st 
(table 3) . Subsiden 

No. 14 during a vis 
8 months after compl 
rk. However, no sub 
the other shafts. I 
nth later and a resu 
in October 1984 indi 
ed 23 feet. It woul 
1 material to reesta 
shaft No. 14 has con 

completed Nov. 1986 
d require about 150 
. No appreciable mo 
sure devices has bee 
itoring. 



able for a 
ce was 
it to the 
etion of 
sidence -was 
nspection 
rvey of the 
cated that 
d require 
blish 
tinued to 
, indicated 
yd 3 of fill 
vement of 
n noted 



Shaft No. 14 was ba 
established practice. I 
control reference to det 
closing the other 13 sha 
failure probably resulte 
temporary bridge that de 
backfilling process or e 
backfilling. No effort 
trash from any of the sh 
or to determine whether 
bridging. Nor was a bul 
base of the shaft to con 
material and prevent it 
washed into the undergro 



ckfilled to locally 
t originally was to be the 
ermine the success of 
fts. However, its sudden 
d from the failure of a 
veloped during the 
xisted prior to 
had been made to remove 
afts prior to backfilling 
there was pre-existing 
khead established at the 
tain the shaft fill 
from spilling or being 
und mine workings. 



CONCLUSIONS 

Eleven shafts were closed u 
Mines designed inverted pyramid- 
plugs. For comparison with more 
backfilling, two shafts had a re 
cap installed after backfilling 
simply backfilled. Initially, a 
closure methods used in the demo 
appeared to be effective methods 
in the Galena, KS, area. Survey 
compiled over a 3-year period in 
the 11 plugs and 2 caps were sta 



sing the Bureau of 
shaped concrete 
conventional 
inforced-concrete 
and a third was 
11 of the three 
nstration project 
for shaft closure 
data, table 3, 
dicate that only 
ble. 



For the shallow shafts backfilling is probably 
a practical way to eliminate safety hazards. 
Backfilling of deeper shafts can require large 
amounts of fill material, particularly where large 
mine openings exist at the base of the shafts. For 
instance a 90-ft high opening at the base of a 
shaft can require up to 57,000 yds 3 of fill 
material to reach the base of the shaft. Further 
subsidence can continue to be a problem with 
backfill ing. 



For any shafts of any 
concrete caps or pyramid pi 
method selected depends to 
cost of labor. For either 
intensive, due in part, to 
safety harness while workin 
mine shaft's potential coll 
caps required the most time 
spent working over a mine s 
concrete than the plugs for 
shaft. 



depth, either installing 
ugs is effective. The 
a great extent on the 
method, labor is cost 
the restrictions of a 
g within the abandoned 
apse zone. The concrete 

exposure of personnel 
haft and required more 

closing a given size 



The inverted pyramid plug was the preferred 
method for closing the shafts in this demonstration 
project. This method was the easiest closure 
device to install and could easily be done assembly 
line fashion for closing a large number of shafts 
in a relatively short time. The plugs required the 
least amount of work in the immediate vicinity of a 
potentially dangerous abandoned mine shaft. The 
most time intensive installation step was spent in 
setting up the 17-ton crane, while avoiding 
locating it over any known underground workings, to 
safely place the plug forms in the shafts. Most of 
the labor was spent in fabricating the disposable 
forms. However, this was performed in a welding 
shop away from the demonstration site. Evaluation 
of the closure devices is being continued over a 
longer period of time to further prove their 
stabil ity. 

LITERATURE CITED 

Clerc, F. L. 1907. The Ore Deposits of the Joplin 
Region, Mo. Trans. AIME, v. 38, pp. 320-343. 

Crane, W. R. 1901. Methods of Prospecting, Mining 
and Milling in Kansas Lead and Zinc District. 
KS Geol. Surv., v. 8, pp. 177-387. 

Dressel, W. M., M. C. McFarland, and J. C. Brown. 
1986. Post-Mining Hazards of the Kansas- 
Missouri-Oklahoma Tri-State Zinc-Lead Mining 
District. Guidebook to the Geol. and 
Environmental Concerns in the Tri-State 
Lead-Zinc District, MO, KS, OK. Assoc, of 
Missouri Geologists 33rd Annu. Meeting, Joplin 
Mo, Sep 26-27, pp. 47-54. 

Dressel, W. M. and J. S. Volosin. 1985. Inverted 
Pyramid Shaped Plugs for Closing Abandoned 
Mine Shafts-Galena, Ks., Demonstration 
Project. BuMines IC 8998, 14 p. 

Fejes, A. J., R. C. Dyni, J. A. Magers, and L. B. 
Swatek. 1985. Subsidence Information for 
Underground Mines-Literature Assessment and 
Annotated Bibliography. BuMines IC 9007, 
86 p. 

Genie Eng. LTD. 1983. The CLWYD Mine Cap., 

Technical Lit., Unit 16, Garston Industrial 
Estate, Window Lane, Liverpool 19, United 
Kingdom, 1983. 

Gibson, A. M. 1972. Wilderness Bonanza Tri-State 
District of Missouri, Kansas, and Oklahoma. 
Univ. OK Press, 362 pp. 

Haworth, E. 1901. History of Geography, Geology, 
and Metallurgy of Galena-Jopl in Lead and 
Zinc. Kansas Geol . Surv . , 8, pp. 177-387. 

Hay, R. 1893. Geology and Mineral Resources of 
Kansas. KS State Bd. Agri . , Biennial Rep. 8, 
pt. 2, pp. 99-162. 

Henrich, C. 1892-93. Zinc-Blende Mines and Mining 
Near Webb City, Mo. Trans. AIME, v. 21, 
pp. 3-24. 



Luza, K. V. 1983. A Study of Stability Problems 
and Hazard Evaluation of the Oklahoma Portion 
of the Tri-State Mining Area. Contract 
J0100133, OK Geol. Surv. -Univ. OK, BuMines OFR 
76-83, 147 pp. 



412 



Martin, A. J. 1946. Summarized Statistics of 
Production of Lead and Zinc in the Tri -State 
(Missouri-Kansas-Oklahoma) Mining District. 
BuMines IC 7383, 67 p. 

McCauley, J. R., L. L. Brady, and F. W. Wilson. 
1983. A Study of Stability Problems and 
Hazard Evaluation of the Kansas Portion of the 
Tri-State Mining Area. Contract J0100131, KS 
Geol. Surv.--Univ. KS, BuMines OFR 75-83, 
193 pp. 

McFarland, M. C, and J. C. Brown. 1983. Study of 
Stability Problems and Hazard Evaluation of 
the Missouri Portion of the Tri-State Mining 
Area (contract J0100132, MO Dept. Nat. 
Resour., Div. Geol. and Land Surv.). BuMines 
OFR 97-83, 141 pp. 

National Coal Board. 1982. Treatment of Disused 
Mine Shafts and Adits. London, 86 pp. 



Norris, J. D. 1968. AZn: A History of the 

American Zinc Company. State Hist. Soc. WI, 
244 pp. 

Plyn, J. 1904. The Joplin District. Mines and 
Miner., v. 24, Feb., pp. 329-330. 

Stewart, D. R. 1986. A Brief Description of the 
Historical, Ore Production, Mine Pumping, and 
Prospecting Aspects of the Tri-State Zinc-Lead 
District of Missouri, Kansas, and Oklahoma. 
Guidebook to the Geol. and Environmental 
Concerns in the Tri-State Lead-Zinc District, 
MO, KS, OK. Assoc, of Missouri Geologists 
33rd Annu. Meeting, Joplin Mo, Sep 26-27, 
pp. 16-29. 

U.S. Environmental Protection Agency (EPA). 1975. 
Inactive and Abandoned Underground Mines., 
440/9-75-007, 338 p. 



413 

U.S. GOVERNMENT PRINTING OFFICEi 19 8 8-210-296/80271 



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